17

Table 2 Major dose-reconstruction studies at DOE facilities in the United States

Site Object of assessment Reference
Completed
Nevada Test Site Hypothetical individuals Henderson and Smale (1990)
Nget al. (1990)
Bouville et al. (1990)
Collective Henderson and Smale (1992)
Whicker et al. (1996)
Specific individuals Simon et al. (1995)
Till et al. (1995)
Hanford Works Hypothetical individuals Farris et al. (1994a)
Farris et al. (1994b)
In process
Hanford Works Specific individuals Davis et al. (1995)
Fernald Feed Materials
Production Center Miller et al. (1994)
Rocky Flats Plant Miller et al. (1994)
Oak Ridge Site Miller et al. (1994)
"eho vneeriog Labo ratory Miller et al. (1994)
Savannah River Site Miller et al. (1994)



Abstract

In the late 1950s and early 1960s, a series of tests was conducted at or near the Nevada Test
Site to study issues involving plutonium-bearing devices. These tests resulted in the dispersal of
about 5 TBgq of 239.240Py on the surficial soils at the test locations. Access to the sites is strictly
controlled; therefore, it does not constitute a threat to human health at the present time.
However, because the residual 239Pu decays slowly (half-life of 24,110 y), the sites could indeed
represent a long-term hazard if they are not remediated and if institutional controls are Jost. To
investigate the magnitude of the potential health risks for this no-remediation case, we defined
three basic exposure scenarios that could bring individuals in contact with 239.240Py at the sites:
(1) a resident living in a subdivision located at a test site, (2) a resident farmer, and (3) a worker
at a commercial facility. Our screening analyses indicated that doses to organs are dominated by
the internal deposition of Pu via the inhalation pathway, and thus our risk assessment focused on
those factors that affect inhalation exposures and associated doses, including inhalation rates,
activity patterns, tenure at a residence or occupation, indoor/outdoor air relationships, and
resuspension outdoors. Cancer risks were calculated as a function of lifetime cumulative doses
to the key target organs (i.e., bone surface, liver, and lungs) and risk factors for those organs.
Uncertainties in the predicted cancer risks were analyzed using Monte-Carlo simulations of the
probability distributions used to represent assessment parameters. The predicted cancer risks for
the resident farmer were more than a factor of three times higher than the suburban resident at
the median risk level, and about a factor of ten greater than the reference worker at a commercial
facility. At 100 y from the present, the 5, 50, and 95th percentile risks for the resident farmer at
the most contaminated site were 4 x 10-6, 6 x 10-5, and 5 x 10-4, respectively. The principal
sources of uncertainty in the estimated risks were population mobility, the relationship between
indoor and outdoor contaminant Jevels, and the dose and risk factors for bone, liver, and lung.


Attachment 21:

Layton, David, et. al. Risk Assessment of Soil-Based Exposures to Plutonium at
Experimental Sites Located on the Nevada Test Site and Adjoining Areas. California:
Lawrence Livermore National Laboratory. UCRL-ID-113605. June 1993.


DATE ©
FILMED
9/18 1

END


Before preparing a detailed dose assessment of an environmental contaminant, it is often
helpful to prepare a screening-level analysis to identify the exposure pathways that contribute the
most to absorbed doses, and conversely, to eliminate from further consideration those pathways
that contribute little to absorbed doses. Information from the first two analyses is then used to
examine the transport pathways of concern and afterwards, to calculate the doses to key organs.

To relate the internal doses to potential health risks, we next examine dose-response data for
estimating the probability of incurring cancer per unit dose to target organs. We have included in
Figure 2 additional feedback loops between the dose-response assessment, dose assessment, and
exposure assessment tasks to indicate that there is a need to exchange information at each point
in the assessment process. The final task in the methodology is to characterize the cancer risks
resulting from the exposures to 23%240pu-contaminated soils/dusts at the different sites along
with the uncertainties of the estimated risks. To quantify the overall uncertainty in calculated
cancer risks, we use a Monte-Carlo simulation technique to propagate the uncertainties
associated with the 239.240Py source terms (i.e., concentrations in soil), plutonium transport,
human exposures, absorbed doses, and dose-response relationships for target organs. Each of
the following discussions evaluates separately the uncertainties of key assessment parameters
and the concluding section on risk characterization presents the overall results.

2. Residual Plutonium in Soil at Shot Sites

The tests resulting in Pu dispersal were not always conducted in the same manner. A review
(U.S. DOE, 1988) indicates that some device detonations occurred on concrete pads and in
concrete bunkers. For example, Double Tracks and Clean Slate I were conducted by detonating
one or more warheads on concrete pads, while in Clean Slate II and III warheads were exploded
within bunkers. Post-shot cleanup operations also varied among the test series. For the
Operation Roller Coaster tests, remediation efforts included the scraping of soils and other debris
around the ground-zero (GZ) area to a single location where they were covered with clean soil.
Decontamination efforts at the Plutonium Valley sites included the excavation of trenches in
which contaminated soils and other test-related materials were buried, In contrast, there were no
remediation efforts associated with the Project 57 (Area 13) test.

Systematic efforts to characterize the residual 239:240Py contamination of soils at the various
shot sites were completed under the auspices of the Nevada Applied Ecology Group. The
concentrations of 239.240py in soils at the shot sites and estimates of the associated 239.240py
inventories are given in Gilbert (1977), but are based in part on earlier work given in Gilbert and
Eberhart (1974) and Gilben et al. (1975). The plutonium content of soils at each of the sites was
characterized statistically using a stratified, random sampling approach. In this approach the area
to be sampled was subdivided into separate strata or sampling areas in such a way that the
variability of plutonium concentrations in soil samples taken from within a given stratum is
smailer than the variability of the concentrations in samples taken across strata. Strata were
defined for each of the shot sites using data obtained from FIDLER (Field Instrument for the
Determination of Low-Energy Radiation) surveys. This particular field instrument uses a 5-in.
Nal detector to measure 60-keV gamma emissions from the decay of 241Am (432-y half-life),
which is a decay product of 241Pu (14.4-y half-life). Although most of the plutonium in nuclear
weapons occurs as 239Pu, 241Pu is also present. Asa result, 241Am activities in contaminated


11

Beck, H.L., and Anspaugh, L.R. Development of the County Database: Estimates of Exposure
Rates and Times of Arrival of Fallout in the ORERP Phase-II Area. Comparison with
Cumulative Deposition-Density Estimates Bases on Analyses of Retrospective and
Historical Soil Samples. U.S. Department of Energy Nevada Operations Office, Las Vegas,
DOE/NV-320, 1991.

Bouville, A., Dreicer, M., Beck, H.L., Hoecker, W.H., and Wachholz, B.W. Models of
Radioiodine Transport to Populations within the Continental U.S. Health Phys. 59: 659-
668 (1990).

Caldwell, G.G., Kelley, D.B., and Heath, C.W., Jr. Leukemia Among Participants in Military
Maneuvers at a Nuclear Bomb Test. J. Am. Med. Assoc. 244: 1575-1578 (1980).

Caldwell, G.G., Kelley, D., Zack, M., Falk, H., and Heath, C.W., Jr. Mortality and Cancer
Frequency Among Military Nuclear Test (Smoky) Participants, 1957 through 1979. J. Am.
Med. Assoc. 250: 620-624 (1983).

Davis, S., Kopecki, K., Hamilton, T.E., Amundson, B., and Myers, P.A. Hanford Thyroid
Disease Study, Pilot Study Final Report, Fred Hutchinson Cancer Research Center, Seattle,
1995.

Dose Assessment Advisory Group. Final Report. U.S. Department of Energy Nevada
Operations Office, Las Vegas, 1987.

Dunning, G.M. Fallout from Nuclear Tests at the Nevada Test Site. In: U.S. Congress (1959),
Vol. 3; pp. 2021-2053.

Farris, W.T., Napier, B.A., Ikenberry, T.A., Simpson, J.C., and Shipler, D.A. Atmospheric
Pathway Dosimetry Report, 1944-1992. Pacific Northwest Laboratories, Richland, WA,
PNWD-2228 HEDR, 1994a.

Farris, W.T., Napier, B.A. Simpson, J.C., Snyder, S.F., and Shipler, D.A. Columbia River
Pathway Dosimetry Report, 1944-1992. Pacific Northwest Laboratories, Richland, WA,
PNWD-2227 HEDR, 1994b.

Friesen, H.N. A Perspective on Atmospheric Nuclear Tests in Nevada. U.S. Department of
Energy Nevada Operations Office, Las Vegas, DOE/NV-296, 1985.

Glasstone, S., and Dolan, P.J., Eds. The Effects of Nuclear Weapons. Government Printing
Office, Washington, 1977.


The age-dependent dosimetry for 239.240Pu is based on a model that was developed by
Leggett (1985) and later modified for use in ICRP Publication 56 (ICRP, 1990). Our
implementation of the modified model (illustrated schematically in Figure 17) utilizes the age-
specific physiologic and biokinetic parameter values given in Tables 1.-1 and A-1 of
ICRP Publication 56, respectively. In addition, our model includes a modified version of the
ICRP (1979) lung model, as shown in Figure 17. In the respiratory portion of the model, inhaled
239,240Py aerosols (i.e., Class Y aerosols with an AMAD of 1 1m) are deposited on the surfaces
of the nasal passages (denoted NP) and the trachea and bronchial tree (denoted TB) and are
transferred instantaneously to blood and the GI tract via lung subcompartments a, b, c, and d.
The standard ICRP lung model integrates the dose to these subcompartments; however, because
they are cleared rapidly, little accuracy is lost in omitting them. The parameters fa, fp, fc, and fa
indicate the fractions of the deposited 239.240Pu transferred to blood or lungs (see Fig. 17).
Aerosols reaching the pulmonary region that are deposited (determined from the deposition
fraction fp) are transferred to lung e, h, and g subcompartments (the quickly cleared
f subcompartment is omitted in our model), The lung subcompartment h clears to the pulmonary
lymph nodes, while the e and g subcompartments clear to blood and the GI tract, respectively.
Once in the blood, some 239.240py is distributed to the systemic organs shown in Figure 17, but
most is translocated to skeleton and liver. The Leggett model divides the skeleton into cortical
bone and trabecular bone. These bone types are subdivided into three biologically-relevant
compartments for estimating doses: (1) bone surface, (2) bone volume, and (3) bone marrow.
‘The liver is treated as a single compartment in the modified Leggett model.

The portion of the model used dealing with bone dosimetry is based entirely on the bone-
recirculation approach of Leggett (1985) as implemented in ICRP Publication 56. In this
approach, specified fractions of all Pu circulating in blood continuously enter compartments
representing cortical-bone and trabecular-bone surfaces, respectively, and the Pu is then lost
(with first-order kinetics) to the volume and marrow portions of those types of bone. The
cortical and trabecular marrow also receives 239.240py directly from blood, again through first-
order processes. Integrated 239:240Pu a-particle dose to bone is calculated as summarized in
Table 7.4 of ICRP Publication 30 (ICRP, 1979). According to this dosimetry model for a+
emitters in bone, all bone surfaces plus all trabecular (i.e., red) bone marrow comprise the
biologically relevant set of bone-related target organs. Thus integrated dose to bone surfaces
includes (a) dose delivered to that surface from 239.240py situated within cortical and trabecular
surfaces plus (b) the doses delivered from 239.240Py situated within cortical and trabecular bone
volumes, appropriately discounted for the limited range of 239.240Pu a-panticles and the
geometric properties of the source and target compartments involved. Similarly, integrated dose
to red marrow includes 100% of the 239.240Pu dose contained within trabecular marrow plus
those doses (again, appropriately adjusted) delivered to that marrow from trabecular-bone surface
and volume compartments.

Four target organs (liver, lung, bone surface, and red bone marrow) were selected for analysis
based on a preliminary review of dose-response data on those organs. The uptake, translocation,
deposition, and excretion of Pu through the various compartments and corresponding integrated
doses to the target organs was calculated using a system of ordinary differential equations, which
we have solved numerically using Mathematica (Version 2.0, Wolfram, 1991). These equations
included time-dependent parameters, as noted above. The continuous relationships between age

32


soils at GMX. In earlier experimental work at the GMX site, Anspaugh er al. (1976) calculated
resuspension rates for 239.240Pu ranging from 2.7 x 10712 to 4.8 x 10°'0 s-1, As a comparison,
Hartmann et al. (1989) reported a resuspension rate of 9 x 10-1! s~} for pre-Chernobyl 239,2A0Py
at rural sites in Germany, using a contamination depth of 1 cm. Their rate would actually be
lower, though, if a 5-cm depth was used to represent the amount of 239,240py available for
resuspension. Collectively, these studies indicate that the resuspension rates for the NTS shot.
sites probably range between 10-!2 to 10-1! s-1 (3.15 x 10-5 to 3.15 x 104 y-).

The time-dependent change in the concentration of 239.240pu in soil resulting from
resuspension and radioactive decay can be estimated from the initial concentration in soil as

Cs(t) = Cs(0) expl-(Aa+ Ar) 4), @

where t is in years, Ag and Ay are the rate constants for radioactive decay and resuspension (both
rate constants in y-!). If we use a resuspension rate of 5.5 x 10°12 5:1 (or 1.7 x 10-4 y-))(the
mean of the limiting values 10-12 to 10-1! s-1) to represent the net loss of 239:240Py from surface
soils at the shot sites, the environmental half-life of 239.240Pu is 4000 y, or about 17% of the half-
life due to radioactive decay alone.

The value of Ca can be calculated as a function of time from Cs using either the mass-loading
or resuspension-factor approaches. With the mass-loading method the time-dependent
concentration of 239.240py in air becomes

Ca(t) = C(O) expl-Cra+ Ar) f] x TSP x EH 6)
and using a resuspension factor, it is
Cat) = CeO) expla + An) 1x11 x St, (6)

where 1) is a factor that converts the concentration of Pu in soil on a mass basis to an areal basis,
as a function of soil density and depth of contamination. With C. equal to 1 Bq g-!, the value of .
Ca is 3 x 10-5 Bq m-3, assuming that the particulate mass loading at the shot sites is
approximately 3 x 105 g m3 (the average of the two values reported in Shinn er al., 1986) and
the enhancement factor is one. For a nominal soil density of 1.5 g cm-3 and a contamination
depth of 5 cm, the value of 7) is 7.5 x 104 g m~2, By setting Eq. 5 equal to Eq. 6, St can then be
calculated as (TSP x EF)/n. With TSP and EH values noted above, Sr equals 4 x 10°10 ml, a
result that is comparable to the resuspension factors reported above.

To quantify the uncertainties in the estimates of Cg at the different shot sites, we will
propagate the uncertainties associated with the individual parameters in Eq. 5, using a Monte-
Carlo technique (Crystal Ball, v. 2, Market Engineering Corp., 1991), The means and standard
errors of the C; values for the shot sites are represented in Table 1. To avoid the generation of
negative concentration values from the Monte-Carlo sampling of the normal distributions for Cs,
- we converted the mean and standard errors to their lognormal equivalents (i.c., geometric mean
and geometric standard error) using the equations given in Ng er al, (1990). Resuspension rates
are characterized using a triangular distribution with minimum and maximum values of

14


4. Transport Processes for Plutonium
at Shot Sites

In order to estimate potential inhalation exposures to residual 239.240Py at the shot sites, we must
determine the concentrations of Pu in both indoor and outdoor air. People spend routinely more
than 90% of their time indoors, and hence differences between the concentrations of 239-240Pu in
indoor and outdoor air will affect directly estimated inhalation exposures. While outdoors an
individual can inhale soil particles containing 239.240Pu that have been resuspended into air by
wind action. Plutonium-contaminated soil can reach the indoor environment via three principal
mechanisms: (1) infiltration of airborne soil particles through the shell of a house, (2) inflow
with makeup air drawn into a building by a ventilation system, and (3) tracking of contaminated
soil onto floors, followed by resuspension into indoor air by human aztivities. Suspended
particles in indoor air can be removed by filters that are part of air-handl.ng systems in homes
and commercial establishments. In the following subsections we estiinate concentrations of
239,240Py in outdoor air at the shot sites and then determine the levels indoors for the buildings
associated with the residential, farm, and commercial exposure scenarios.

4.1. Concentrations of Pu in Surficial Soils and Outdoor Air

The primary loss mechanism of 239240Pu in surficial soils in the dry environs of the NTS is
aeolian erosion. In locations where there are no human activities such as vehicular treffic over
unpaved roads, excavation activities, etc., the suspension of soil particles is caused by the
movement of wind over the land surface. Most of the early work on this transport mechanism for
soil contaminants focused on the resuspension of radioactive substances deposited onto surface
soils after nuclear testing. There are many environmental and contaminant-specific factors that
affect the transfer of a soil contaminant from soil to air, and the complex nature of this transfer
process has made it difficult to develop analytical approaches that provide accurate predictions of
resuspension (see Nicholson, 1988; Sehmel, 1980; and Smith et al,, 1982; for reviews of
resuspension models). As an altemative, empirical approaches have been developed to estimate
the average concentration of a soil contaminant in air. One commonly used method involves the
calculation of a resuspension factor, which is the ratio of a contaminant's concentration in air to
its concentration in soil expressed on an area basis (¢.g., Bq or ig m-3 + Bq or pg m2), The
resuspension factor (denoted S,) for a substance initially introduced to a surficial soil as a result
of atmospheric deposition or aerial application decreases with time as the material undergoes
various weathering processes. Anspaugh et al. (1975) proposed this time-dependent relationship
for the resuspension factor

S¢(t)=104 xexp(-0.15Vi)m! +109 m7), (i)

where ¢ is time in days since initial addition to soil. This relationship was based on data from
several sites measured at various times post deposition. The first term in Eq. 1 dominates during
the early years following a contaminating event to reflect weathering processes, but diminishes
with time to the second term, which Anspaugh et al. set equal to the resuspension factor

2


Cin _ foMgAgfaR + Qr(1-Ep)fm TSP EH

Cs (Anvg + QrE,(1- fm) + Qrfm}

(for a commercial facility). (11)

The model parameters can be broken down into two basic groups, one dealing with
movement of particles and the other building properties. Particulate-related parameters include
the mass loading of soils/dust on floors, the resuspension rate, the settling rate of particles, and
the penetration factor. Building properties include the surface area of floors, the air-exchange
rate, building volume, the air-recirculation rate, and the particle-removal efficiency of filters in
the air-handling system. Many of these parameters are best characterized as statistical
distributions, while others can be considered as constants for the purposes of the exposure
assessment. In the following discussion we describe the parameter distributions and values for
the residential, farm, and commercial buildings.

The loading of 239.240Pu on floors (Lf) can be estimated as the product of the concentration
of 239.240py in tracked-in soil, the fraction of indoor dust that is derived from outdoor soil, and
the loading of soil/dust on floors. Fergusson ef al. (1986) estimated that between 45 to 50% of
the dust in a house is derived from soil, based on analyses they completed of the elemental
compositions of soil and house dust. More recently, Calabrese and Stanek (1992) calculated that
31% of the dust indoors had an outdoor origin, and Allott e¢ al. (1992) reported that
concentrations of !37Cs in indoor floor dust were between 18 and 48% of the 137Cs
concentrations in outdoor soil. Based on these literature values, we have adopted a uniform
distribution between 0.2 and 0.5 (a range that encompasses the literature values) to represent the
random values of fo for the three types of buildings. The loadings of soil/dust on floors reported
in the literature vary from 0.136 to 0.870 g m~2 (see Appendix A). The geometric mean of the
mass loadings is 0.42 g m-? with a geometric standard deviation of 1.88. The resuspension rate
of particles from floors to indoor air depends largely on human activities, such as walking,
playing, etc., that result in the mechanical disturbance of floor surfaces. Healy (1971) estimated
a resuspension rate indoors of 5 x 10~“ h-!, using a mix of vigorous, active, moderate, and quiet
domestic activities. Murphy and Yocom (1986), in an analysis of the migration of particles into
the indoor environment, concluded that the resuspension rate in houses was about 10-4 h-!. We
will use a uniform distribution between 10“ and 10-3 to reflect a range of human activities that
resuspend particles from floor surfaces.

Removal of a particulate contaminant by gravitational settling onto floor surfaces is governed
by its concentration in air and the average settling velocity of the contaminated particles. The
deposition of fine particles onto wall and ceiling surfaces due to diffusion and other mechanisms
is estimated using a deposition velocity. Settling/deposition velocities for fine (i.e., < 2.5 jim) to
coarse (2.5 to 15 fim) particles suspended in indoor air range from about 0.18 to 25 mh7! (i.e,
0.005 to 0.7 cm s~}) (Sinclair et al., 1992). Diemel et al. (1981) made concurrent measurements
of airborne lead in homes and the rate of mass accumulation on horizontal collection plates, and
from their data we estimate a settling velocity of lead of 1.2m h-!, which is an intermediate
value for fine and coarse particles. 1n contrast, we estimated that the average settling velocity for
dust particles in that study was 2.7 m h~!, Raunemaa et al. (1989) reported a deposition velocity
of 9.2 mh”! for total coarse particles indoors and 4.9 m h- for total fine particles. Deposition
velocities for individual elements can differ greatly. For the crustal element potassium,
Raunemaa ef al. reported deposition velocities of 15 and 8.8 m h-! for coarse and fine particles,

18


Indoors at home Indoors away from home

Outdoors at home Outdoors away from home

Figure 8. Four basic categorizations of indoor and outdoor locations for use In assessments of
exposures to soll-derlved contaminants at given location.

Indoors at home

Fraction of day

Outside at home

0 10 20 30 40 50 60 70
Age, years

Figure 10. Fractions of time Individuals spend indoors at home and outside at home (date are
from Wiley, 1991; Wiley et a/., 1991).

22


throughout the Basin and Range province of Nevada and is therefore likely to occur at the NTS
as well. The other three land uses are more dependent on the availability of ground water. The
Tesident rancher, for example, would use well water to support beef cattle and dairy cows as well
as a family garden. Residential and commercial developments could obtain potable water from
one or more wells supplying multiple users.

The conceptualization of the basic exposure pathways that could bring individuals in contact
with the 239.240py derived from soils at the shot sites is shown in Figure 7. Exposure pathways
include ingestion of garden vegetables, consumption of animal products derived from
contaminated soils, soil ingestion, and inhalation of airborne particles. Ingestion and inhalation
exposures result in the uptake of 239.240py and subsequent transfer to body organs, which are
then irradiated by the emission of alpha particles during the radioactive decay of 239,240Py, An
integrated exposure and dose model that addresses the pathways shown in Figure 7 was
developed specifically for the NTS by the Nevada Applied Ecology Group (NAEG) (see Martin
and Bloom, 1980). Kercher and Anspaugh (1991) have performed subsequently a series of
sensitivity and uncertainty analyses of the NAEG model. They found that the inhalation pathway
produces the greatest doses to individual organs. For example, inhalation of 239.240py-
contaminated particles resuspended from soils produced 100% of the dose to the upper-
respiratory tract, lungs, and lymph nodes; and 94.9% of the dose to the liver, kidneys, and bone.
On a whole-body basis, inhalation represented 94.8% of the total dose and ingestion 5.2%. The
principal reason for the difference between the two routes of exposure is the limited uptake of
239,240Py across the GI tract: Plutonium in oxide form has a very low solubility, and since only
dissolved species are effectively transported across the epithelial lining of the intestines, the
biouptake of Pu is small. After Pu is deposited in the lungs, though, it is slowly translocated to
other systemic organs, resulting in higher organ doses than for ingestion exposures.

The analyses that Kercher and Anspaugh performed of the sensitivity of organ doses (i.e.,
lung, liver, bone, and kidney) to changes in model parameters indicated that doses were most
sensitive to changes in the concentration of 239.240py in soil, the mass-loading factor used to
estimate the resuspension of 239.240Pu from soil to air, the respiration rate, and dose factors that
convert the activity in an organ to the amount of energy deposited. Other less important
parameters dealt mainly with biophysical factors such as the fraction of particles deposited in the
lungs, clearance rates, and mass of target organs. These screening-level analyses indicate that
the inhalation pathway should be the primary focus of the risk assessment for 239.240py in soils at
the NTS. Consequently, our reference land-use scenarios will only consist of residential
(suburban) housing, resident farming, and commercial development, cattle grazing is excluded
on the basis that the ingestion of 239:240Py-contaminated meat products would not constitute a
significant radiological health risk.

10


21

Table 6. Methods of dose reconstruction using the analysis of environmental residues

¢ Deposition densities, historical or current data

Short-lived radionuclides (historical data only)
ms

137¢§

Uranium

239+240py

239+240py, plus the ratio of 240Py-to-239Pu

e Historical measurements of extemal gamma-exposure rate



Introduction ..........06+.

7.
Appendix A.
References...

Contents

Methodology Overview ........:..0+0
Residual Plutonium in Soil at Shot Site:

Conceptualization of Potential Exposure Pathways and Screening Analyse:
Transport Processes for Plutonium at Shot Sites
41. Concentrations of Pu in Surficial Soils and Outdoor Air ......ssssssscsesseocseeecees 12
4,2. Concentrations of Pu in Indoor Air

Human Factors Describing Exposure Scenarios..
5.1, Activity Patterns .......

5.2. Age-dependent inhalation rates ..

5.3. Residential and Occupational Mobility .
Organ-—Specific Doses and Risk Factors.

6.1. Organ—Specific Doses from Inhalation Exposure:
6.2. Organ-Specific Risk Factors
Predicted Cancer Risks and Associated Uncertainties



Cin(t), Cou(t) = time-varying concentration of 239.240Py in indoor or outdoor air, Bq m-3;
BRin. BRom= age-dependent breathing rate while indoors or outdoors, m3 d-!;

Tins Tout= age-dependent amount of time spent indoors or outdoors at a given
location, d y~};

pce?) = cumulative lifetime dose to age 70 y for the ith organ for inhalation
exposure at time t, uGy Bq-!.

One difficulty with the formulation of Eq. 19 is that it cannot be used easily to simulate the
residential and occupational mobility of individuals, that is, the lengths of time spent at a given
location as a function of age (as defined by the integration limits). A simpler formulation can be
derived by treating the indoor/outdoor concentrations of 239-240Py as constants over exposure
periods and by assuming that the residence times of individuals at a given location are distributed
randomly with age. The revised formulation then becomes

DY = M[CouFi, out + CinFi, in] (20)
where
b
Fi out = oo JBRout(t) Tourlt) DC? «t) dt, (21)
a
b
= 1 70
Fin = bra ! BRin(t) Tig (t) DC? Ct) dt, and (22)
where

Methe residence time at a given residential, farm, or occupational
location, y; and

F;, in and Fj, out = €xposure-normalized dose-conversion factors for indoor and outdoor
inhalation exposures over the period defined by the integration limits
aand b, m3-Gy y-!_Bq-],

For workplace exposures to Pu aerosols, the parameters Tou and Tin are treated as constants
(ie., 5.4 and 84 d y-}, respectively) and the breathing rate for active hours is used as the effective
inhalation rate for both indoor and outdoor locations.

31


The lower part of Table 2 indicates dose-reconstruction activities at DOE sites that are
currently in progress. As part of an epidemiologic study at Hanford, thyroid doses will be
calculated for specific individuals. In addition, a variety of different dose reconstructions are now
underway at a number of sites: The Fernald Feed Materials Production Center, the Rocky Flats
Plant, the Oak Ridge Site, the Idaho National Engineering Laboratory, and the Savannah River
Site. It is important to consider whether or not it is worthwhile to continue to undertake so many
dose-reconstruction activities in lieu of lessons learned from those studies already completed.

Table 3 indicates some of the reasons that dose-reconstruction studies might be
undertaken. Certainly, the Nevada Test Site was a prime example where there was a known large
release from which the possibility of biologic effects could be presumed. The Hanford dose
Teconstruction falls into the next category, where there was a stunning release of formerly
classified information that indicated substantial releases, and there was also at least one deliberate
release with public exposure. That situation led to the next category where, at Hanford and
elsewhere, issues of social justice arose. Some members of the public felt very strongly that they
had been wronged and, in some cases, harmed. In such cases, there are certainly reasons to
undertake some kind of a dose-reconstruction study and, perhaps, an epidemiologic study.

Almost all of the studies indicated in the lower part of Table 2 were, in fact, undertaken
for social justice reasons. Whether or not this is an adequate reason for conducting such studies,
which typically cost a few tens of millions of dollars, is something that should be considered.
Certainly, a significant fraction of the public thinks social justice concerns are a valid reason, and
they believe that it is reasonable for the government to pay for a preferably independent study in
return for the government having allowed a situation to develop where the public has come to
question even the good will of its government. Scientists may disagree with the need or the
usefulness of such studies, but the desire of the public for more information and reassurance that
they are not at significant risk has perhaps been the compelling reason for the implementation of
many dose-reconstruction studies in the U.S. Of course, there are other reasons for dose-
reconstruction studies to be undertaken, including a desire to advance knowledge of radiation-risk
factors, particularly for low dose radiation received at low-to-moderate dose rates, and some


at exposure and cumulative dose to age 70 y to target organs per unit respiratory intake ( pen
in Gy Bq-!) are illustrated in Figure 18. The cumulative doses to the four organs decrease with
age at exposure as the period of integration decreases. Nonmonotonic changes in the shape of
the curves are due to corresponding age-dependent changes in organ weights, organ-specific
transfer fractions, and excretion rates. The curves are in the form of interpolating functions for
cumulative doses determined at various ages of exposure. Our cumulative dose estimates for
each of the target organs are all within a few percent of the inhalation dose coefficients (in Sv
Bq"!) for 239Pu given in ICRP Publication 56 for six ages when exposure is assumed to occur
{ie., 3 mo, ly, 5 y, 10y, 15 y, and adult (> 17 y)).

For the residential and farm exposure scenarios, the integration period (a - b) encompasses the
ages of 0 to 70 y, whereas for a commercial land use the integration period encompasses the age
span of the worker population (i.¢., 19 to 65 y). Table 3 contains the average dose-conversion
factors for the two types of domestic exposure scenarios as well as the commercial exposure
scenario. For a unit concentration of 239.240Py in aie and exposure period, bone surfaces receive
the greatest doses, followed by the lungs, liver, and bone marrow. The bone marrow receives
less than 10% of the dose to the other organs. Differences between the F,;, and F;., variables
are due to breathing rates used for indoor/outdoor locations and the amounts of time spent
indoors and outdoors, The values of Fou for a resident farmer would increase by a factor of
approximately five if the value of Tow were equivalent to 3 h d-! (12.5% of a day) spent outside
in the immediate vicinity of a farmhouse. An increased breathing rate associated with a
metabolic equivalent energy expenditure of 1.8 would only increase the active breathing rate by
about 25%.

Bone surface

Lifetime dose, pGy/Bq

h_ Bone marrow

Cy) 10 20 30 40 50 60 70
Age, y

4. A

Figure 18. Cumulative lifetime doses to age 70 y for selected body organs based on
Instantaneous respiratory exposures at different ages.


te

Skeleton

Blood

Other Soft
Tissue

Modified Legget-ICRP dosimetry model for inhaled and/or ingested plutonium. C = cortical, T =

= surface, V = volume, e-j = ICRP lung compartments, f (A) = fraction
(or Ist-order exponential, indicated by
f radiation dose without mass transfer.

Figure 17.
trabecular, R = Red, M = marrow, s
deposited via (or loss constant for) processes presumed instant
thinnest arrows and typeface). Dashed arrow signifies a deposition o

See text for further details.



Table 4. Summary of the geometric standard deviations of the 3940py burdens (i.e,

activity/unit mass of tissue) in autopsy samples of livers, lungs, and bone tissue analyzed by

various researchers.
Cee mal

Geometric standard

Organ Sample size deviation Ref.
Liver 171 (male) 2.03 Griffith and Guilmette (1991)
Liver 139 (female) 191 Griffith and Guilmette (1993)
Liver 10 193 Kawamura and Tanaka (1983)
Liver 59 25 Takizawa et al. (1987)
Liver 104 25 Singh et al. (1983)
Liver to> 23 Singh et al. (1983)
Liver 238 2.04 Popplewell et al. (1985)
Lung 10? 15 Singh et al. (1983)
Lung 10 19 Singh et al. (1983)
Lung, 1o 1.48 Kawamura and Tanaka (1983)
Lung 66 21 Takizawa et al. (1987)
Lung, 2B 266 Popplewell ef al. (1985)
Lung 25 148 Irlweck ef al. (1980)
Bone a 13 Singh et al. (1983)
Bone wb 7 Singh et al. (1983)
Vertebrae 16 20 Takizawa et al, (1987)
Femur B 1.63 Popplewell et al. (1985)
Vertebrae 6 230 Kawamura and Tanaka (1983)

2 Autopsy samptes from Washington, DC.
> Autopsy samples from Grand Junction, CO.

6.2. Organ-Specific Risk Factors

Several major reviews have been published recently concerning the effectiveness of various
forms of radiation to induce cancer in humans; several of these studies are of direct interest here.
Broad general reviews have been published by the United Nations Scientific Committee on the
Effects of Atomic Radiation (UNSCEAR, 1988), the U.S. National Academy of Sciences’
Committee on the Biological Effects of Ionizing Radiation (BEIR) (USNAS, 1990), and the
International Commission on Radiological Protection (ICRP, 1991). Other reviews, which
specifically consider the effects of alpha radiation, are of more specific interest. These include
ICRP Publication 50 (ICRP, 1987), which considers the lung-cancer risk from exposure to radon
decay products, and a prior report from the USNAS (1988), which also considered the risk from
radon and other internally deposited alpha—emitting radionuclides, specifically including Pu.

It is well established that most injected Pu ends up in the skeleton and liver (ICRP, 1986),
and the latest ICRP metabolic model contains the assumption that of the total amount of Pu

36


Attachment 20:

Boice, John D., et. al. “A Comprehensive Dose Reconstruction Methodology for Former
Rocketdyne/Atomics International Radiation Workers.” Health Physics. Vol. 90, No.5,
May 2006. 409-430.


18

Table 3. Reasons why dose-reconstruction studies might be undertaken

e Known large release that can be presumed to have a biologic effect
¢ Stunning revelation of formerly classified data
« Operational releases
¢ Deliberate releases with public exposure
Social justice
© The public believes it has been harmed.
© The public believes it has been wronged.
To advance knowledge of risk factors
e “Because it’s there.”



24

Table 9. Characteristics of the releases from the Nevada Test Site

Characteristic Value
Number of atmospheric tests >100
Approximate total yield 1 x 105 tonnes
Approximate energy release? 1 x 10!5 cal
Fission-product atoms created? 3 x 1026
Cesium-137 released? 6 x 1015 Bq
lodine-131 released? 6x 1018 Bq
Collective external dose? 1 x 104 Gy
Collective thyroid dose* 1x 105 Gy

These values follow from the 1 Mt total yield and conversions provided in
Glasstone and Dolan (1977).

bFrom Henderson and Smale (1992)

°From Whicker et al. (1996)


20

Table 5. Methods of performing dosimetry on individuals themselves

Chromosome-translocation analysis of circulating lymphocytes
Electron paramagnetic analysis of teeth

Whole body counting for long-lived radionuclides

Analysis of material excreted in urine or feces

Analysis of tissues collected at autopsy or exhumation



COPYRIGHTED MATERIAL

PLEASE REFER TO CITED JOURNAL
TO VIEW ARTICLE


15

US. Congress. Low-Level Radiation Effects on Health. Hearings before the Subcommittee on
Oversight and Investigations of the Committee on Interstate and Foreign Commerce,
House of Representatives. Government Printing Office, Washington, Serial No. 96-129,
1979,

U.S. Congress. “The Forgotten Guinea Pigs.” A Report on Health Effects of Low-Level
Radiation Sustained as a Result of the Nuclear Weapons Testing Program Conducted by
the United States Government. Report Prepared for the Use of the Committee on Interstate
and Foreign Commerce, United States House of Representatives and Its Subcommittee on
Oversight and Investigations. Government Printing Office, Washington, Committee Print
96-IFC 53, 1980.

Whicker, F.W., Kirchner, T.B., Breshears, D.D., and Otis, M.D. Estimation of Radionuclide
Ingestion: The “PATHWAY” Food-Chain Model. Health Phys. 59: 645-657 (1990).

Whicker, F.W., Kirchner, T.B., Anspaugh, L.R.,, and Ng, Y.C. Ingestion of Nevada Test Site
Fallout: Internal Dose Estimates. Health Phys. 71: 477-486 (1996).

WHO. Environmental Health Criteria 25. Selected Radionuclides: Tritium, Carbon-14,
Krypton-835, Strontium-90, Iodine, Caesium-137, Radon, Plutonium. World Health
Organization, Geneva, 1983.


Puskin, J.S. (1992), “An Analysis of the Uncertainties in Estimates of Radon-Induced Lung
Cancer, Risk Anal. 12; 277-285.

Puskin, J.S., N.S. Nelson, and C.B. Nelson (1992), “Bone Cancer Risk Estimates,” Health Phys.
63, 579-580.

Raunemaa, T., M. Kulmala, H. Saari, M. Olin and M.H. Kulmala (1989), “Indoor Air Aerosol
Model—Transport Indoors and Deposition of Fine and Coarse Particles,” Aerosol Sci.
Technol. 11, 11-25.

Riumallo, J.A., D. Schoeller, G. Barrera, V, Gattas, and R. Uauy (1989), “Energy Expenditure in

Underweight Free-Living Adults: Impact of Energy Supplementation as Determined by
Doubly Labeled Water and Indirect Calorimetry,” Am. J. Clin. Nutr. 49, 239-246.

Sehmel, G.A. (1980), “Particle Resuspension: A Review,” Environ. Int. 4, 107-127.

Shah, J.J., R.L. Johnson, E.K. Heyerdahl, and J.J. Huntzicker (1986), ““Carbonaceous Aerosol at
Urban and Rural Sites in the United States,” J. Air Poll. Control Assoc. 36, 254-257.

Shinn, J.H., E.H. Essington, F.L. Miller, T.P. O'Farrell, J.A. Orcutt, E.M. Romney, J.W. Shugart
and E.R. Sorom (1989), “Results of a Cleanup and Treatment Test at the Nevada Test Site:
Evaluation of Vacuum Removal of Pu-Contaminated Soil,” Health Phys. 57, 771-779.

Shinn, J.H., D.N. Homan and C.B. Hoffmann (1986), A Summary of Plutonium Aerosol Studies:
Resuspension at the Nevada Test Site, Lawrence Livermore National Laboratory, Livermore,
CA, UCRL-90746.

Sinclair, J.D., L.A. Psota-Kelty and C.J. Weschler (1985), “Indoor/Outdsor Concentrations and
Indoor Surface Accumulations of Ionic Substances,” Atmos. Environ. 19, 315-323.

Sinclair, J.D., L.A. Psota-Kelty, G.A. Peins, and A.O. Ibidunni (1992), “Indoor/Outdoor
Relationships of Airborne Ionic Substances: Comparisons of Electronic Equipment Room
and Factory Environments,” Atmos. Environ. 26A, 871-882.

Singh, N.P., M.E. Wrenn, and S.A. Ibrahim (1983), “Plutonium Concentration in Human
Tissues: Comparison toThorium,” Health Phys. 44, Suppl. 1, 469-476.

Smith, W.J., F.W. Whicker and H.R. Meyer (1982), “Review and Categorization of Saltation,
Suspension, and Resuspension Models,” Nucl. Safety 23, 685-699.

Solomon, R.L. and J.W. Hartford (1976), “Lead and Cadmium in Dusts and Soils in a Small
Urban Community,” Environ. Sci. Techol. 10, 773-777.

Stannard, J.N. (1988), Radioactivity and Health, A History, Pacific Northwest Laboratory,
Hanford, WA, DE88013791, DOE/RL/018300-TS59.

Takizawa, Y., A. Hisamatsu, and T. Abe (1987), “Concentration of Fallout Plutonium in Tissues
of Japanese who Died during 1980-1984,” Radiat. Res. 109, 245-255.

UNSCEAR, Sources, Effects and Risks of lonizing Radiation. 1988 Report to the General
Assembly, with annexes, (United Nations Scientific Committee on the Effects of Atomic
Radiation, New York, 1988).

US Bureau of the Census (1991), American Housing Survey for the United States in 1989, U.S.
Government Printing Office, Washington, DC, Current Housing Reports, H- 150/89.

50


Table 3. Exposure-normalized dose-conversion factors calculated for different exposure

scenarios and organs.
Exposure-normalized dose-conversion factors
Fin Fw
Exposure scenario/tissue m3 - uGy yt - Bq74
Residential/Farm
Bone surfaces 6.7 x 104 39x18
Liver 1.6 x 104 9.4.x 102
Lung 4.1 x 104 27x 108
Bone marrow 38x10 23x10
Commercial
Bone surfaces 3.1x 104 20x 103
Liver 67x10 43x 102
Lung 1.7x 104 11x10
Bone marrow 17x18 11x10?

Variability in the F,;, and Fj, parameters between individuals will be due to
imerindividual variations in breathing rates, time spent indoor/outdoors, and biodosimetric
parameters such as the fractions of inhaled Pu deposited in different tissues, residence times in
organs, etc. As @ means of characterizing the variability of the exposure-weighted dose-
conversion factors, we reviewed data from various studies that included measurements of the
burdens of 239.240Pu in autopsy samples of livers, lungs, and bone from people exposed to fallout
Pu. Table 4 summarizes the geometric standard deviations of the tissue analyses. Most of the
researchers found that burdens of 239.240py in organs were lognormally distributed, with GSD
values ranging from 1.48 to 2.66. We have chosen a GSD value of 2 to represent the variability
in our dose-conversion factors for thos: organs, as this represents an intermediate value of the
GSDs listed in Table 4, Because the 239.240Py burdens in organs are unlikely to be independent,
we must also quantify correlations between the doses estimated for the different organs. Bunz]
and Kracke (1983), for example, reported positive correlations between 239.240Py in 30 autopsy
samples of livers and lungs (based on the log-transformed burdens) from individuals who lived
in Germany. However, neither they nor Takizawa er al. (1987) found a statistically significant
correlation between age and burden. To explore further the relationship between the 239,240py
burdens in lungs, liver, and bone, we calculated the linear correlation coefficients between
239,240Py burdens in autopsied organs from 28 individuals who had been exposed to 239.240py
fallout in Great Br‘tain (Popplewell er al., 1985). The linear correlation coefficient between the
log-transformed 239:240Pu burdens in lung and liver was 0.88, between lung and bone (femur) it
was 0.77, and between bone and liver it was 0.81. We will use these correlations when
generating correlated random variables representing the Fj in and Fj, gut parameters in the
Monte-Carlo simulations.

35


2400 m3/h-! (based on a typical air-handling rate of 1400 ft? min-!), If the unit operates an
average of 30% per year (denoted fon), the effective recirculation rate is 1200 m3? h-!. The
particle-removal efficiency (Ef) of standard fibrous filters in domestic HVACs is approximately
0.5 (Lefcoe and Inculet, 1975). Air recirculation rates in the commercial facility are assumed to
be for an office environment, with Q, defined as the product of the building volume (V) and an
air exchange rate (Ach), defined as a uniform distribution between 10 and 20 air-exchanges per
hour. The resulting limits of the uniform distribution of Q, for the commercial building volume
of 2200 m3 (i.e., 930 m2 x 2.4 m) are 2.2 x 104 and 4.4 x 104 m3h-!. The outdoor air makeup
(fm) fraction is defined as a uniform distribution between 0.1 to 0.2 (from Hayes, 1989; Weschler
et al., 1983). The filter efficiency for coarse particles is also assumed to be 0.5. Table 2
summarizes the values used to represent the various input parameters to Eqs. 10 and 11.

Monte-Carlo simulations of Eq. 10 using the appropriate parameter values and distributions
for a suburban house (with central air conditioning and heating) resulted in a lognormal
distribution of the Cin/C; ratios with a GM of 6.8 x 10-6 g m-? and a GSD of 1.89. The GM and
GSD of the Cjn/C; ratios for the farm house (no central air conditioning or heating) were
9.1 10-6 g m-3 and 1.95, respectively. The larger ratio for the farm house was due primarily to
the lack of indoor filtration of suspended particles by a central air conditioning/heating unit. The
commercial facility, which has a continuously operating air ventilation and filtration system,
achieved the lowest ratio of Cin/Cs, with a GM of 5.3 x 10-6 g m-3 and a GSD of 1.56. The
geometric means of the ratios of Cin/Ca (equivalent to the indoor/outdoor concentration ratios of

Table 2, Summary of the parameter values used to determine the concentrations of 239,240Py
in the indoor air of houses and commercial buildings.

eel

Residential Fann Commercial
Variable Units _Dstribution Statistics
P unitess Uniform Min. (0.2) Max. (0.3) Min. (0.2) Max. (0.3) Min. (0.2) Max. (0.3)
Ach wl Lognormal GM (0.67)GSD (2.0) GM @.67)GSD(2.0) GM (0.67) GSD (2.0)
‘pg mil Triangular Mode (6) Min. (2) Made (6) Min. (2) and = Mode (6) Min. (2) and
and Max. (10) Max. (10) Max. (10)
Aq m2 Constant 150 150 930
fy unitess Constant 0.75 0.75 0.75
fo unitless Uniform Min, (0.2) Max. (0.5) Min. (0.2) Max. (0.5) Min. (0.2) Max. (0.5)
Mél g m2 Lognormal GM (0.42) GSD (1.88) GM (0.42) GSD (1.88) GM (0.42) GSD (1.88)
h m Constant 24 24 24
Q m3h-1 — Constant 2400 - -
a m3h-1 Uniform - - Min. (22,000)
Max. (44,000)
fon unidess Constant 03 - -
Ef unidess Constant 0.5 - 0.5
TSP pg m~3 Lognormal GM (28) GSD (1.6) GM (28) GSD (1.6) GM (28) GSD (1.6)
EH unidess Constant 1 1 1
R wl Uniform 104 w 103 104 103 104 t0 10-3

IEEE

20


upon earlier studies conducted by the U.S. Department of Energy’s Environmental Measurements
Laboratory (Krey and Beck 1981) and depended upon a variety of historical data and on the
collection of soil samples for the analysis of environmental residues. For the contemporary soil
samples, the analysis consisted of measuring !37Cs, 239+240py and the ratio: 240Pu-to-239Pu.
The problem was that most of the 137Cs at these locations actually came from global weapons
fallout, not from the Nevada Test Site itself. Consequently, a much more elaborate procedure
was required to determine how much of it came from the Nevada Test Site. Cesium-137 itself is
not of particular interest, but all of the short-lived radionuclides that came with it delivered most
of the dose. However, if one knows how much of the currently present !37Cs at a particular
location came from a particular test, one can infer the presence of all of the short-lived
radionuclides (Hicks 1990; Beck and Anspaugh 1991).

Table 9 presents a summary of the NTS-release characteristics. There were more than one
hundred atmospheric tests. The cumulative yield was one megaton, which is not too much
considering that one Soviet test, by itself, was 58 megatons. However, this one very large test
was conducted at Novaya Zemlya, a very isolated location. The largest test performed by the
United States was about 15 megatons, but this was carried out in the Pacific, not in Nevada. The
137C¢ released at the NTS was 6 PBq, or 6 x 10!5 Bq. That is about ten percent of the amount
that was released by the Chernobyl accident (UNSCEAR 1988). Six EBgq, or 6 x 10!8 Bg, of 1311
were released, which is about 20 times more than what was released from Chemobyl.

Fortunately, most of the !31I released from the NTS was injected high into the troposphere,
where it was extensively diluted and had time to decay before much of it came to earth.

A summary of the local population exposure in Phases I and Il is that the collective
external dose was about 104 Gy, and the collective thyroid dose was about 10° Gy. Absorbed
doses from the ingestion of fallout-contaminated food were also calculated for 21 organs in
addition to the thyroid (Ng et al. 1990; Kirchner et al. 1996, Whicker et al. 1996). These results
are summarized in Fig. 4, where are plotted the collective doses to the indicated organs. The
highest collective dose is to the thyroid, and the internal dose to this organ is the only one for


ye 1261
re Olé

~
8
Date

1,400,000
1,200,000
1,000,000 -+
800,
600.

Ak/sayng,

Fig. 6. Yearly releases of radionuclides to the Columbia River from the Hanford Works (from
Farris et al. 1994).


20 T T T T Tv T T
15 oa 4
2 meee eee eet
e 4
i 10 a 7
£ - palette et eee
= ~
3 4
O 5 ----- Breathing rate during active hours = _|
7 Breathing rate indoors at home
f —--— Breathing rate at rest 4
0 L l | ! n i
0 10 20 30 40 50 60 70
Age, y

Figure 14. Age-dependent changes In the breathing rates at rest, Indoors at home, and during
active hours.

frequency for the prescribed location, but also the duration of the exposure. To estimate the
length of time spent at residences in different urbanizations (i.e, rural suburban and farm), we
have analyzed statistical data from a national survey of U.S. housing in 1989 (U.S, Bureau of the
Census, 1991). Figure 15 presents empirical probability distributions of the lengths of time that
householders reside at domiciles situated in suburban areas or at farms in rural parts of the U.S.
A substantial fraction (i.e., 0.45) of the suburban householders sampled in 1989 spent less than
four years at their current residence. The mean residence time was about ten years. As might be
expected, those living on farms were far less mobile. They resided an average of 20 years in
their homes — twice the mean residence time for those suburbanites living in rural areas. The
number of years that employees might work at an unspecified commercial facility can be
estimated from survey data on workers from different occupations in the U.S. (U.S. Bureau of
the Census, 1991). In 1987 the median occupational tenures for 27 major occupations (defined
as those with 650,000 workers or more) ranged from 1.9 to 21.) y (excluding occupations that
are not associated with a single work location, such as truck drivers and carpenters), The
geometric mean tenure of the occupations was 6 years and the geometric standard deviation was
1.74 (see Figure 16). These occupational tenures will tend to overestimate the number of years
spent at a given commercial location because an individual may change work locations.

27


100

§ E
8 L
~ L
2
2
5
2
w 10 7 f
4 E : |
Z i
Ss |
Qa r fi i
3 Poa j H
9 | i H |
fe) t ri | ' H i
i
1 L it i
01.1. 1-5 10 2080 50 70809095 99 99.999.99

Percent

Figure 16. Logprobability plot of the median occupational tenures for 27 major, nonfarm
occupations in 1987 (data trom U.S. Bureau of the Census, 1991). The geometric mean [s 6 y and
the geometric standard deviation Is 1.74.

6. Organ-Specific Doses and Risk Factors

The cancer risk associated with the inhalation of 239,240py in environmental media is
governed by the doses to various organs and the risk factor (i.é., the incremental probability of
cancer per unit dose) for each organ. Accordingly, we have developed estimates of organ-
specific doses and risks for the exposure pathways discussed in the previous section. The basic
objective of the dose assessment is to determine the cumulative doses of 239,240Py to critical
organs in individuals as a function of the concentrations of 239,240py in environmental media,
human factors (¢.g., land use, inhalation rates, activity patterns, mobility, and diet), and the age-
dependent retention, translocation, and excretion of plutonium.

29


The cancer risk of exposures to 239-240Py can be expressed as

a
R= YDP xR <9)
izl

where

R = the incremental risk of incurring cancer, unitless;

bp? = the cumulative lifetime dose to age 70 y for the ith organ of n critical organs, Gy;
and

Rj = the linear risk factor for the ith organ, Gy~).

With this methodology, cancer risk is calculated directly from the quantities of energy
deposited in key organs resulting from the decay of retained 239240Py and the dose-response
relationship between deposited energy and the probability of cancer for those organs. In contrast
with this approach, ICRP (1991) methodology relies on the determination of equivalent doses to
individual organs (i.e., doses that are adjusted to account for the biological effectiveness of
different radiations, in units of sieverts, Sv) and the use of tissue-weighting factors, which reflect
the different probabilities of developing cancer in various organs of a body uniformly irradiated
by external gamma radiation (based largely on cancer-mortality data for A-Bomb survivors). In
the sections below we present our estimates of organ-specific doses resulting from the inhalation
exposures to 239,240py derived from contaminated soils associated with the alternative land uses
discussed earlier, and we also assess risk factors for estimating the probability of cancer for
internally irradiated organs.

6.1. Organ-Specific Doses from Inhalation Exposures

Doses to specific organs are calculated from the exposure conditions used to define the
frequency and duration of an individual's inhalation of airborne 59.240Py and its concentrations
in the indoor and outdoor environments. The lifetime cumulative dose to age 70 y for a given
organ resulting from an inhalation exposure over a specified period of time (defined by the
integration limits a and b) can be calculated from

b b
DP? = J[Cour6t) BRour(t) Tout) DC7(t) dt + [Cin(t) BRin(t) Tin(t Dc?%(t) dt (19)
a a
where
Dp? = lifetime cumulative dose to age 70 y for the ith organ resulting from

inhalation exposures to 23%240Py at a given location over an exposure
period defined by the integration limits a and b, wGy;

30


‘(1661 yBnedsuy

Pure yoog WO) azg Isa] ePRASN ayy OJ APNys UONDNUJsuOIas-aSOp ay} JOy suOIBas APUG “¢ “BL

ASEA

-s,

A BIL

‘A RAPID

A AD

‘SCOTTSBLYFF

MEMPH

ao

CORPUS CHRISTI °

[3] Phase-i Region
Phase-H Region
A Phase-tt Sites

_NTS



A survey of the predicted cancer risks for the various shot sites given in Table 5 indicates that
the highest risks are associated with Area 11 or Plutonium Valley. The resident farmer's cancer
risk, for example, exceeds 10“ at the 95th cumulative percentile tevel calculated for an assumed
loss of institutional control at 100 years from the present. At 50,000 and 100,000 y in the future,
the predicted cancer risks are all below 10-6, which indicates that we have covered the principal
time domain during which future adverse effects might occur.

To gain additional insights to the sources of uncertainty in the cancer risks, we analyzed how
much of the variance in our risk estimates was derived from each of the following parameters:
resuspension rates, concentrations of 239.240Py in soil, indoor/outdoor concentrations of
239,240Py, mobility, and organ-specific risk factors for cancer. For this analysis, we completed a
series of Monte-Carlo simulations in which only one parameter was allowed to vary randomly,
while the other parameters were held constant at their mean value. Because the calculated risks
are determined essentially from a multiplicative model, the sum of the individual log variances in
the various input parameters should equal (approximately) the log variance in cancer risk. The
fraction of total variance in estimated risk that is attributable to a given parameter is determined
simply as the ratio of the log variance in cancer risk due solely to the variation in a given
parameter and the total log variance in risk from all the random variations in the model
parameters. In Figure 19 we show the contributions to total variance in the predicted cancer risks
for the residential scenario for Area 13 at 100 y from the present. The principal source of
variance in cancer risk is the mobility of residents (68.6%), followed by the exposure-weighted
dose factors (17.1%), the ratio Cin/C, used to relate levels of 239:240pu in outdoor soil to
concentrations in indoor air (9.4%), and the cancer-risk factors (4.9%). Variation in the
resuspension rate and the concentration of 239240Pu in soil contributed little to the total variance.
However, at 50,000 y the resuspension rate becomes the principal source of variation in predicted
risks. This is because the resuspension rate appears within an exponential term as the negative
product of time (see Eq. 5), and so as time increases, the value of the resulting exponential term
eventually becomes the primary determinant of the concentration of 239.240py in air and hence
cancer risk as well. Thus, for any given exposure scenario occurring over a person's lifetime, the
Tesuspension rate will contribute little to overall risk, but when risks are predicted far into the
future, the resuspension rate can become an important parameter.

40


used extensively in the United States. Whole-body counting for some materials that remain in the
body for a long time, such as Sr, is another useful method, as is the analysis of tissues at
autopsy or exhumation. Such methods have found widespread application in Russia, and

Dr. Degteva will discuss these methods further.

Some of the methods of dose reconstruction involving the analysis of environmental
residues are indicated in Table 6; these types of analyses have been very useful in studies in the
United States. In cases where there was a large-scale release, such as that at the Nevada Test
Site, there are elements in the environment that can still be measured. These include radionuclides
such as Sr, 137Cs, and 239+240py. A lot can also be learned about the source of plutonium and
cesium by measuring the isotopic ratio: 240Py-to-239Pu (Krey and Beck 1981; Beck and
Anspaugh 1991). The Feed Materials Production Center at Fernald, Ohio, was mainly a uranium-
processing center. As uranium is a very long-lived radionuclide, the collection and analysis of soil
samples has been useful in the reconstruction of material released at that site (Stevenson and
Hardy 1993). Iodine-129 is another long-lived radionuclide (half-life of 16,000,000 years) and is
sometimes a good tracer for past deposits of its short-lived, but more damaging sibling, 13'1. The
usefilness of current measurements of 1291 in inferring the doses to the thyroid from 131] has been
demonstrated for the Chemobyl accident (Straume et al. 1996). In addition, having access to
both historical and current measurements of external gamma-exposure rate is very useful, as has
been the case for both the Techa River study in Russia and the Nevada Test Site study in the U.S.
There is a variety of other measurements that can be made even at times long after the major
exposure has occurred, because residues are still in the environment. If the relative abundance of
the source term is known, in fact, the deposition of material can be completely reconstructed
based upon the measurement of even a single radionuclide (Anspaugh and Daniels 1996).

NEVADA TEST SITE DOSE RECONSTRUCTION
The off-site radiation-review (ORERP) project at the Nevada Test Site was a very large,

multi-person, multi-agency, modular study. The names of the Principal Investigators and other
key individuals, along with their organizations, are shown in Table 7. As the NTS study was the


researchers tremendously enjoy the challenge of dose reconstruction. Dose reconstruction is a
very difficult activity and sometimes, perhaps, scientists seek to undertake projects because of the
challenge and their own interest.

Table 4 indicates some of the methods of dose reconstruction. The methods are arranged
hierarchically, with those producing the most credible results at the top. Certainly, if it were
possible to examine and extract a sample from every individual of concer, and if such a sample
could function as a dosimeter, a more credible dose assessment would result. The next most
desirable method of dose reconstruction would be dosimetry based on materials that could be
taken from people's homes. A good example of this would be extracting quartz grains from
bricks, and then doing a thermoluminescent-dose measurement using the quartz as a dosimeter.
This, in fact, has been done in association with dose-reconstruction work for the Japanese
survivors of the atomic bombings (Maruyama et al. 1987) and for persons living downwind of the
Nevada Test Site (Haskell et al. 1995). It can also be very useful to analyze environmental
residues, as such residues may provide a reliable indication of the current or past presence of
radioactive materials. Sometimes, if the process that resulted in the exposure occurred a long,
time ago and if there are no environmental residues, there may be no choice other than to try to
reconstruct the releases from process information and to then infer doses using an atmospheric
transport model followed by models of the movement of radionuclides through food chains to
man and metabolism in man. Sometimes, releases can only be inferred from second-hand
information. If only a collective dose is sought, calculations of atmospheric transport are not
needed. The lowest order in the hierarchy indicated in Table 4 is simply an inference of releases
coupled to the use of global dose factors, such as those published by the IAEA (1985), the
UNSCEAR (1982), or the WHO (1983).

Table 5 indicates some of the individual biological dosimeters. It is possible to extract a
blood sample and do an analysis of stable translocations within chromosomes of circulating
lymphocytes (Lucas et al. 1995). This method has a sensitivity of about 100 mGy. Electron
paramagnetic analysis of teeth is another more recently developed possibility (Haskell et al. 1995),

and this method is currently receiving very wide interest in Russia and Ukraine. It has not been


16

Table 1. Schedule for Section IT on Dose Reconstruction

Presenter Time Topic
Lynn Anspaugh 20 min. Introduction and Overview of Dose Reconstruction;
Lessons Learned from Studies in the United States
Marina Degteva 25min. Reconstruction of Radiation Doses in Populations near the
Techa River

Alt 15 min. _ Open Questions/Discussion on Dose Reconstruction



if it was possible to reconstruct intemal doses. Eventually internal doses were reconstructed (Ng
et al. 1990; Whicker et al. 1990, 1996; Kirchner et al. 1996), and the process was carried out in
the public arena under the guidance of an academic and political advisory group (DAAG 1987).
But it is of historical interest that the commitment was made to perform this study, even though at
the time it was not certain that the project could actually be carried out. Several features about
the project were unique: One was the determination to make the maximum use of the available
measurements, rather than to rely on models; another was to calculate the best estimate of dose
along with an expression of uncertainty, this led to the first major implementation of stochastic
models for dose reconstruction; another first was to carry out the study under the full awareness
of the public through a federal advisory group that included representatives from the governors of
four affected states (Nevada, Utah, Arizona, and California); and finally, the study was conducted
by a large number of investigators from national laboratories, universities, government agencies
including the U.S. Environmental Protection Agency, and private companies. The latter forced
the development of a modular approach.

The most important historical measurements that could be used for the dose
reconstruction were those of external gamma-exposure rate; such measurements were made after
each individual event at the NTS. Fig. 2 indicates the summation of these measurements and their
extrapolation with time to provide an overall idea of the extent of the external gamma doses in the
nearby communities. These historical estimates of external gamma exposure were confirmed by
the later studies of the ORERP. After the ORERP had begun and achieved some initial successes,
it became clear that the area of consideration in Fig. 2 was not sufficient to include the majority of
people who had received exposures of concern. Thus, consideration was given to broadening the
study to include the Phase II region indicated in Fig. 3. Possible expansion into the Phase II area
(essentially the remainder of the U.S.) indicated in Fig. 3 was considered. Based upon the
analysis of soil samples taken in the areas marked in Fig. 3, however, this expansion was
considered by the Dose Assessment Advisory Group not to be necessary (DAAG 1987).

As there had not been comprehensive measurements of external gamma-exposure rate in
the Phase II area, a new method of inferring dose was required. The method used was based


US Bureau of the Census (1991), Statistical Abstract of the United States 1991, 111th ed.,
Superintendent of Documents, U. S. Government Printing Office, Washington, DC.

US Department of Fnergy (1988), CERCLA Preliminary Assessment of DOE's Nevada
Operations Office Nuclear Weapons Testing Areas, Vol. 1, US Department of Energy,
Nevada Operations Office, Las Vegas, NV.

USNAS, Health Risks of Radon and Other Internally Deposited Alpha-Emitters (BEIR IV), (U.S.
National Academy of Sciences, Washington, 1988).

USNAS, Health Effects of Exposure to Low Leve's of lonizing Radiation (BEIR V), (U.S.
National Academy of Sciences, Washington, 1990).

Voelz, G.L. and J.N.P. Lawrence (1991), “A 42-y Medical Follow-Up of Manhattan Project
Plutonium Workers,” Health Phys. 61: 181-190.

Weschler, C.J., S.P. Kelty, and J.E. Lingousky (1983), “The "ffect of Building Fan Operation on
Indoor-Outdoor Dust Relationships,” J. Air Poll. Contr. Assoc. 33, 624-629.

Wiley, J.A. (1991), Study of Children’s Activity Patterns, Research Division, California Air
Resources Board, Sacramento, CA.

Wiley, J.A., J.P. Robinson, T. Piazza, K. Garrett, K. Cirksena, Y.T. Cheng, and G. Martin
(1991), Activity Patterns of California Residents, Research Divisivn, California Air
Resources Board, Sacramento, CA.

Wolfram, S. (1991), Ma:hematica, Addison-Wesley Co., Redwood City, CA.

51


Appendix A
Supporting Data

43


References

Allott, R.W., M. Kelly, and C.N. Hewitt (1992), “Behavior of Urban Dust Contaminated by
Chernoby! Fallout: Environmental Half-Lives and Transfer Coefficients,” Environ. Sci.
Technol. 26, 2142-2147.

Anspaugh, L.R. and P.L. Phelps (1974), “Results and Data Analysis: Resuspension Element
Status Report,” in The Dynamics of Plutonium in Desert Environments, Nevada Applied
Ecology Group Progress Report, P.B. Dunaway and M.G. White, Eds., U.S. Atomic Energy
Commission, Nevada Operations Office, NVO-142, 265-297.

Anspaugh, L.R., P.L. Phelps, N.C. Kennedy, J.H. Shinn and J.M. Reichman (1976),
“Experimental Studies on the Resuspension of Piutonium from Aged Sources at the Nevada
Test Site,” in Atmosphere-Surface Exchange of Particulate and Gaseous Pollutants,
Proceedings of a Symposium held at Richland, Washington, September 4-6, 1974, Technical
Information Center, Energy Research and Development Administration, Washington, DC,
pp. 727-743.

Anspaugh, L.R., J.H. Shinn and D.W. Wilson (1974), “Evaluation of the Resuspension Pathway
Toward Protective Guidelines for Soil Contamination with Radioactvity,” in Population
Dose Evaluation and Standards for Man and His Environment, International Atomic Energy
Agency, Vienna, pp. 513-524.

Anspaugh, L.R., 3.H. Shinn, P.L. Phelps and N.C. Kennedy (1975), “Resuspension and
Redistribution of Plutonium in Soils,” Health Phys. 29, 571-582.

Bair, W.J. and W.K. Sinclair (1992), “Response to Drs. Puskin and Nelson Note,” Health Phys.
63, 590.

Bunzl, K. and W. Kracke (1983), “Fallout 239/240py and 238Pu in Human Tissues from the
Federal Republic of Germany,” Health Phys. 44, Suppl. 1, 441-449.

Calabrese, E.J. and E.J. Stanek (1992), “What Proportion of Household Dust is Derived from
Outdoor Soil?” J. Soil Contamin. 1, 253-263.

Clark, C.S., R.L. Bornschein, P. Succop, $.S. Que Hee, P.B. Hammond and B. Peace (1985),
“Condition and Type of Housing as an Indicator of Potential Environmental Lead Exposure
and Pediatric Blood Lead Levels,” Environ. Res. 38, 46-53.

Cohen, A.F. and B.L. Cohen (1980), “Protection from Being Indoors Against Inhalation of
Suspended Particulate Matter of Outdoor Origin,” Atmos. Environ. 14, 183-184.

Davies, D.J.A., I. Thorton, J.M. Watt, E.B. Culbard, P.G. Harvey, H.T. Delves, J.C. Sherlock,
G.A. Smart, J.F.A. Thomas and M.J. Quinn (1990), “Lead Intake and Blood Lead in Two-
Year-Old U.K. Urban Children,” Sci. Total Environ. 9, 13-29.

Diemel, J.A.L., B. Brunekreef, J.S.M. Boleij, K. Biersteker and S.J. Veenstra (1981), “The
Amhem Lead Study 1. Indoor Pollution, and Indoor/Outdoor Relationships,” Environ. Res.
25, 449-456.

47


2

Table 7. Key individuals and organizations who conducted the Nevada Test Site (NTS) Off-Site

Radiation-Review Project (ORERP)

Names

Organization

L. Anspaugh, Y. Ng,? H. Hicks®
B. Church, R. Nutley, D. Wheeler, M. Page®

W. Whicker, T. Kirchner, M. Otis D. Breshears
R. Henderson, R. Smale

H. Beck

-C. Thompson, F. Miller, R. McArthur

G. Quinn, C. Steadman

F. Grossman
B. Maza, T. Mehas, M. Demarre

Lawrence Livermore National Laboratory

U.S. Department of Energy, Nevada
Operations Office

Colorado State University

Los Alamos National Laboratory

U.S. Department of Energy, Environmental
Measurements Laboratory

University of Nevada, Desert Research
Institute

U.S. National Oceanic and Atmospheric
Administration

US. Environmental Protection Agency

Reynolds Electrical and Engineering Co.

®Deceased


Attachment 19:

Anspaugh, Lynn R. Introduction to Section II and Overview of Dose Reconstruction:
Lessons Learned from Studies in the U.S. California: Lawrence Livermore
National Laboratory. UCRL-JC-126357. Jan. 1997.


Engelmann, R.J. (1992), “Sheltering Effectiveness Against Plutonium Provided by Buildings,”
Atmos, Environ. 26A, 2037-22044.

Fergusson, J.E., E.A. Forbes, R.J. Schroeder and D.E. Ryan (1986), “The Elemental
Composition and Sources of House Dust and Street Dust,” Sci. Total Environ. 50, 217-221.

Fergusson, J.E. ard R.J. Schroeder (1985), “Lead in House Dust of Christchurch, New Zealand:
Sampling, Levels and Sources,” Sci. Total Environ. 46, 61-72.

Gilbert, R.O. (1977), “Revised Total Amounts of 239.240Py in Surface Soil at Safety-Shot Sites,”
in Transuranics in Desert Ecosystems, M.G. White, P.B. Dunaway and D.L. Wireman, Eds.,
U.S. Department of Energy, Nevada Operations Office, Las Vegas, NV, NVO-181,
pp. 423-429.

Gilbert, R.O., L.L. Eberhardt, E.B. Fowler, E.M. Romney, E.H. Essington, and J.E. Kinnear
(1975), “Statistical Analysis of 239240py and 241Am Contamination of Soil and Vegetation
on NAEG Study Sites,” in The Radioecology of Plutonium and other Transuranics in Desert
Environments, U.S. Energy Research and Development Administration, Nevada Operations
Office, Las Vegas, NV (NVO-153).

Gilbert, R.O. and L.L. Eberhardt (1974), “Statistical Analysis of Pu in Soils at the Nevada Test
Site—Some Results,” in The Dynamics of Plutonium in Desert Environments, P.B. Dunaway
and M.G. White, Eds., U.S. Atomic Energy Commission, Nevada Operations Office, Las
Vegas, NV, NVO-142.

Griffith, W.C. and R.A. Guilmette (1991), “Multiparameter Analysis of Fall-Out Plutonium
Burdens in Human Liver,” Radiat. Prot, Dosim. 38, 113-119.

Hartmann, G., C. Thom and K. Bachmann (1989), “Sources for Pu in Near Surface Air,” Health
Phys. 56, 55-69.

Hayes, S.R. (1989), “Estimating the Effect of Being Indoors on Total Personal Exposure to
Outdoor Air Pollution,” J. Air Poll. Control Assoc. 39, 1453-1461.

Healy, J.W. (1971), Surface Contamination: Decision Levels, Los Alamos Scientific Laboratory,
Los Alamos, NM, LA-4558-MS.

ICRP (1979), Limits for Intakes of Radionuclides by Workers, \CRP Publication 30 Part 1,
Pergamon Press, New York, NY.

ICRP (1986), “The Metabolism of Plutonium and Related Elements,” Annals ICRP 16 (2/3).

ICRP (1987), “Lung Cancer Risk from Indoor Exposures to Radon Daughters,” Annals
ICRP 17 (1).

ICRP (1990), Age-Dependent Doses to Members of the Public from Intake of Radionuclides:
Part 1, 1CRP Publication 56, Pergamon Press, New York, NY.

ICRP (1991), “1990 Recommendations of the International Commission on Radiological
Protection,” Annals ICRP 21 (1-3).

Irlweck, K., C. Friedmann, and T. Schonfeld (1980), “Plutonium in the Lungs of Austrian
Residents,” Health Phys. 39, 95-99.

48


<=z =~ Clean Slate!

Net counts per minute of 241Am
Strata 1 [2 <5,000
2 s@mTM 5,000-10,000
3 10,000-25,000
Feet 4 MM >25,000

Clean Slate 1! Net counts per minute of 24tAm
? Strata 1 [2 <5,000
Ground zero 2 mmm 5,000-10,000

3 =I 10,000-25,000
4 Mm >25,000

Figure 3. Areal extent of the contamination of 241 am in soils at Clean State! and Il. The
measured 241m Is used as a surrogate for 229,240py (adapted from Gilbert et al., 1975).


The Hanford Works also released significant quantities of material to the Columbia River,
and Fig. 6 shows the releases by year. Most of the radionuclides of concern are short-lived and
resulted from neutron activation of materials dissolved in the cooling water, which was once-
through water from the River. In general doses from this pathway are substantially less than from
the atmospheric pathway (Farris et al. 1994b).

Dose-reconstruction studies at other U.S. sites are still underway, and results are not yet
available. However, in comparison to the releases from the NTS, it seems unlikely that any of the
other sites indicated in Table 2 would likely have been large enough to have produced
epidemiologically detectable health effects.

The results of the U.S. studies might be summarized as follows:

© The dose from virtually any release can be reconstructed (given enough money and time) ,
and used as part of an epidemiologic study,

© The calculated doses may be very uncertain,

© thas been very difficult to detect any biological effects from the large releases from the
NTS,

© It is not likely that health effects will be seen from the !31] releases at the Hanford Works,
based on a comparison to the NTS results, and

The scientific value of future studies at U.S. sites is questionable, although the studies will

likely continue in order to satisfy social justice.
CONCLUSIONS

The dose-reconstruction efforts for the NTS and the Hanford Works have been
completed, and the results are generally well accepted by the scientific community and the public.
Any resulting health effects from these exposures have been very difficult to detect by
epidemiologic study of the exposed population at the NTS; results for the Hanford thyroid study


which the dose is higher than is the dose from external exposure. Other organs of relatively high
dose are those associated with the gastrointestinal tract and the bone (surface and marrow).

Health effects from the exposures of the general population have been studied by a
number of groups, more notably in recent times by the University of Utah. In general, it has been
extraordinarily difficult to actually detect any biological effects. There is not a detectable increase
in leukemia in the general population; but a statistically significant increase has been reported for a
defined subgroup of acute leukemias discovered from 1952-1963 among individuals younger than
20 at the time of exposure (Stevens et al. 1990). (This association is based upon five cases of
leukemia.) The occurrence of thyroid cancer in children was also studied (Kerber et al. 1993). A
statistically significant positive dose-response trend was reported for neoplasms, but not for
thyroid nodules or for carcinomas. Thus, whatever health effects have occurred, they are not
easily detected by means of epidemiologic study. One significant divergence of the experiences at
the NTS and Chernobyl is the high rate of childhood-thyroid cancer that has been observed
following the Chernobyl accident (Likhtarev et al. 1994). Based upon published data, it has been
estimated that the collective thyroid dose from Chemobyl is 106 Gy (Likhtarev et al. 1996); this is
about ten times higher than the collective dose to the thyroid from the NTS, even though the
release from the NTS was considerably higher. The higher dose from the Chernobyl release is due
to the higher population density and the injection of the !311 at a lower altitude.

THE HANFORD WORKS AND OTHER DOSE-RECONSTRUCTION STUDIES

The major effort of the Hanford Environmental Dose Reconstruction (HEDR) has been
completed (Farris et al. 1994a,b). The major release of radionuclides was to the atmosphere
during the early days of the operation. The monthly releases of !3!] are shown graphically in
Fig. 5. The total release of 1311 is estimated to have been 30 PBq (Heeb 1994), which is about
200 times less than the release from the NTS. As the location at Hanford was also relatively
isolated during the mid 1940s, it seems unlikely that any biologic effect might be found here in
view of the difficulty in finding any effect as a result of the releases from the NTS. However, a
thyroid-health-effects study is underway (Davis et al. 1995), and the results are yet to be reported.


respectively. They estimated that the average mass—median aerodynamic diameters (MMAD)
for all deposited particles was 5 jum. These data suggest that gravitational settling is a more
important removal mechanism than deposition, and hence we assume the floor surface is the
primary surface of dust accumulation. Settling velocities are represented as a triangular
distribution with a mode at 6 m h7! and lower and upper limits of 2 and 10 m h-!,

The median surface area of floors (Aq) in U.S. houses is about 150 m? (U.S. Bureau of the
Census, 1991), and if we assume that most ceilings are 2.4 m high (8 ft), then the associated
volume (V) is 360 m3. These values will be used to represent both the farm and suburban
dwellings. For this assessment we will assume that the commercial facility has 930 m? of floor
space (i.e., 10,000 ft 2). Since only a portion of floors are actually accessible for walking,
playing, etc., we need to adjust the value of Aq downward for the buildings to obtain an effective
floor area that contributes to resuspension. We estimate that the fraction of floor space available
for resuspension (fa) is 0.75. Ventilation rates of houses vary according to climatic as well as
building characteristics. Nazaroff et al. (1985) estimated that the geometric mean air-exchange
rate (Ach) was 0,68 h-!, with a GSD of 2.01. The penetration factor, P, can be estimated as the
ratio of the indoor-air concentration of a contaminant to its outdoor concentration, provided that
there are no significant sources or tosses (e.g., filtration by heating and cooling equipment as
well as deposition onto interior surfaces) indoors. Penetration factors for contaminants contained
in airborne particles generally fall between 0.2 and 0.5, with the highest values associated with
substances enriched on fine particles. Cohen and Cohen (1980), for example, reported that the
average indoor/outdoor ratios for Ca, Fe, Pb, and Br in homes (which did not have indoor
sources of these contaminants) were 0.23, 0.25, 0.53, and 0.38. These penetration estimates
correspond directly with the MMAD for the four elements presented in a review by Milford and
Davidson (1985) (i.e., 4.64, 3.42, 0.55, and 0.89 jm, respectively), The penetration factor can be
estimated using the following power-law fit to the data:

P=0.4 6-037 (72 = 0.96), (12)

where d is the MMAD in micrometers. Shinn ef al. (1989) reported that the activity-median
aerodynamic diameter (AMAD) of 239.240Py sampled in the ambient air in Plutonium Valley was
5 bm with a GSD of 8.5. Anspaugh and Phelps (1974) determined an AMAD of 3 pm for
airbome particulates collected by high-volume cascade impactors operating at the GMX site.
Thus, we would expect that only about 20 to 30% of the resuspended Pu would penetrate a
house. In our simulations we will characterize the variation in the AMAD values for
resuspended 239,240Py at the sites using a uniform distribution with minimum and maximum
values of 3 and 5 xm.

Airbome particles indoors that are derived from infiltrating air and resuspension from floors
can be removed by filters in heating, ventilation, and air-conditioning (HVAC) systems. For the
purposes of our assessment, we will assume that the suburban residence is equipped with a
central heating/cooling system that recirculates air within the building. The farmhouse, though,
is not assumed to have a central heating/cooling unit, and as a result, there is no internal filtration
of particles. The air-recirculation rate (Q,) for a standard heating/cooling unit (i.e., the heating
component has a heat input of 80,000 Btu h-!) in a suburban residence is approximately

19


e @ Clean
', Double Slate
7 Tracks |i,

t
1
1 Test Range 1
t
1
1

4

t

i)

'

Nellis '

Alt Force Base |
Area 11® | ange Complex |
t

1

1

'

4d

rose

Las Vegas

Figure 1. Locations of experiments at the Nevada Test Site and adjoining areas where plutonium:
bearing devices were detonated.


Table A-1, Summary of the residual concentrations and inventories of 239240Py in soils at
shot sites at the NTS (from Gilbert, 1977).

Survey Strata Area Mean Concentration Estimated Inventory

Site strata m? wi m=? SE) a GSE)
13 1 1,245,000 19 + 034 24 + 0.42
2 2,847,000 58 +t 14 15 + 36
3 108,00 3 + 43 25 + 046
4 74,000 4 t a8 40 + 065
5 19,000 wo 6t (9 210 t (036
6 24,000 820 t (340 2 t 82
Total 4,017,000 6 + 99
s 1 111,300 31 + 0.66 035 + 0073
(GMX) 2 8,400 42 + of 035 + 0.076
3 800 270 +t & 0.22 + 0051
4 1,000 530 + 150 0.53 + O15
5 3.800 46 4 16 002 + 0,006
Total 125,300 15 + 019 150 + «(019
Double 1 176,000 67 + 35 12 + 0.62
Tracks 2 1,600 350 + 250 056 «6+ (0.40
3 800 190 t 59 0.15 + 0.047
4 600 46 + 16 17° «+ (0.60
Total 179,000 36 + (095
Clean 1 157,000 3 + 70 24 + Id
Slate I 2 10,000 6 ot 2 064 + 022
3 8,400 m0 + (35 092 + 0,29
4 1,700 w0 06+ (39 0.20 + 0.066
Total 177,100 420 + 12
Clean 1 351,000 41 +4 13 14 * 046
Slate 2 82,300 72 + 30 60 + 25
3 26,200 270 + «99 71 + 26
4 11,000 260 + «65 29° + 0.72
Total 470,500 7 + 37
Clean 1 1,615,000 12 +t 22 19.4 + 36
Slate HI 2 61,000 5B Ot (6 35 + 0.98
40,000 200 2 + «683 a4 +t 25
4 16,000 370+ «190 59 + 30
Total 1,732,000 37 + 54


PPT ITT Ts
PiPUTi Lapa?
Time

+

ThE

{ow/5) pestary LEt-eulpo)

g°

100000
goooo +
0000 +
70000 +-

Fig. 5. Monthly releases of !3!] to the atmosphere from the Hanford Works (from Farris et al.
1994a).


<=" __ Clean Slate Ill

Fence.

Net counts per minute of 247Am
Strata 1 [— <5000

2 mmm 5000-10,000

3 f= 10,000-25,000

4 mmm >25,000
° Seale 600 ° Seale
75
tT === DoubleTracks ==
N Feet Feet

Fence

“

Net counts per minute of 241Am
Strata 1 = <5000
2 @mam 5000-10,000
3 = 10,000-25,000
4 Gm >25,000

Figure 4. Areal extent of the contamination of 241Am In soils at Clean Slate Il! and Double Tracks.
The measured 24am Is used as a surrogate tor 23°.240py (adapted trom Gilbert et ai., 1975).


“ann ween 7

{ Tea

seal

s6uny:

$8150.

[ seesoui

[_apsn }

_f use

sjeuelp'

eM J9pp

sue

[_ seve:

Keup'

481

‘BO Tero,

HEM UDEWIOIS

ounseyut [fe

EW UO pe;

Ad-ueBs0 ‘esop 8013901109

Organ

Fig. 4. Collective dose by organ from the ingestion of contaminated food (based on data from

Whicker et al. 1996).


Metabolic equivatent
(ratio of total energy intake to BMR)

yasewee oe Josiit ranew

=
i"
r
c

0 10 20 30 40 50 60 70 80
Age, years

Figure 13, Average energy Intakes for males and females expressed as metabolic equivalents
(from Layton, 1993).

and a polynomial fit to the data. The breathing rates for sleep/rest, active hours, and indoors at
home are depicted in Figure 14. Farmers are expected to have higher average energy
expenditures than members of the general public and hence higher breathing rates as well.
Riumallo et al. (1989), for example, reported that the average energy expenditure (in metabolic
equivalents) for six males engaged in light agricultural tasks was about 1.8, which is about 20 to
30% higher than the average values for both males and females depicted in Figure 13. The
increased inhalation for a farmer may also be associated with an increase in mouth breathing,
which would lead to a larger deposition of particles in the lungs. As a default in our exposure
and dose estimates, we will use the age-dependent inhalation values for the general population
and estimate the increased exposures for the case of the resident farmer.

5.3, Residential and Occupational Mobility

The cumulative inhalation exposures to airborne 239.240Py at a given residential or
commercial location are not only a function of an individual's breathing rate and exposure

26


first major dose-reconstruction study to be undertaken in the U.S., it is of interest to review why
this initial study was started.’ An underlying reason was the controversy in the mid seventies
over reports that low doses of radiation might be responsible for reputed increases in the
incidence of cancer among workers at the Hanford Works (Mancuso et al. 1977) and at the
Portsmouth Naval Shipyard (Najarian and Colton 1978). Some more specific reasons are
indicated in Table 8; one of the more compelling was that scientists at the Centers for Disease
Control and Prevention found a case of leukemia among military personnel present at the NTS
during the Smoky nuclear test in 1957, and this sparked the conduct of a larger epidemiologic
study (Caldwell et al. 1980, 1983). Another important event was the publication of a study by
Lyon et al. (1979) that was widely viewed as indicating that there had been a radiogenic increase
in leukemia children exposed to fallout. These reports and other concems led to Congressional
Hearings (e.g., U.S. Congress 1979), demands for the release of all information, and even a
Congressional Report with the provocative title of “The Forgotten Guinea Pigs” (U.S. Congress
1980), the cover of which is reproduced in Fig. 1. In addition thousands of claims were filed
against the government and eventually two major law suits were tried. These actions renewed old
concerns about the correctness of prior estimates of external doses (Dunning 1959) and drew
attention to the fact that there had never been an evaluation of internal doses. This latter concern
had surfaced before, particularly concerning the dose to the thyroids of infants from the ingestion
of 131] with food (Knapp 1963; Mays 1963; Reiss 1963). The U.S. government eventually
Tesponded with a commitment to conduct a study, although the initial commitment made on
March 28, 1979, was only to collect and disseminate data on fallout and possible health effects
(DAAG 1987). A second commitment followed a few months later and was to reconstruct, “...in
so far as possible...,” estimates of exposures and doses to the offsite people. The commitment to
conduct the study resulted in the formation of the ORERP, under the management of the U.S.
Department of Energy’s Nevada Operations Office.

At the time of the commitment for this dose-reconstruction study, external doses had been
estimated and published (e.g., Dunning 1959; historical results are also reviewed extensively by
Anspaugh et al. 1990), but had not been extensively peer reviewed. It was not known at that time

+ Further perspective is provided in Anspaugh (in press).


19

Table 4. Methods of dose reconstruction

© Individual biologic analysis
¢ Chromosome-translocation analysis of circulating lymphocytes
¢ Electron paramagnetic analysis of teeth
¢ Dosimetry of materials in homes
Analysis of environmental residues
¢ Deposition densities, past or current
e External gamma-exposure rates (past)
¢ Known releases, plus atmospheric models
© Inferred releases, plus atmospheric models
* Known or inferred releases



Table A-1. (Continued)

Survey StrataArea Mean Concentration Estimated Inventory
Site strata m? yCi m-2 (SE) ca GSE)
11 2 8,200 30 + 18 025 + O15
B Site
3 6,00 220 + 35 13 +t 033
4 3,300 1,400 +t 300 4.6 + 09
Total 17,500 62 +11
11 2 16,400 34 + 2 056 + 036
C Site
3 13,300 88 + 3 049 + «(G14
3,500 1,400 + 390 a9 + 14
5 300 6,200 2 2,800 19 + 0846
Total 25,800 78 4°17
11 2 32,300 46 t 95 15 + 031
D Site
3 13,300 220 + 86 29 tl
4 4,900 990 + 370 49 + 18
5 2,900 2,700 + 840 7.8 + 24
Total 53,400 WA + 32
CD Overlap 6 62,200 12 t 5.2 075 + 52

BSL a

Table A-2. Summary of data on lead and dust in the indoor environment.

Number Loadings on floor/carpets,
of Pb in dust
Location dwellings gg? HgPb m2 mg Dust m? References
Christchurch, NZ 65 Bis. 423 804 Fergusson and
Schroeder, 1985
22 590 349 525
Edinburgh, Scotland 10 381 52 136 Laxen et al., 1988
Birmingham, UK 97 336 a 1799 Davies et al., 1990
Cincinnati, OH"
Putlic 2 350 200 571 Clark et al., 1985
Rehabllitated 46 623 370 594
Private 33 1410 750 532
Champaign-Urban, IL 12 500 440 870 Solomon and
Hartford, 1976
Arnhem, The 107 282 123 * 255 Diemel et al., 1981
Netherlands (coarse)
957
(fine)

Excludes deteriorating and dilapidated housing

45


a. House Filter

Filtration
Or fon Et
HVAC unit mn
v
~<—;—  AchV Ach Ca PV «t—4-—— Ca
Exfiltration infiltration Resuspended
particles
Li AnfaA AnVg
Resuspension Gravitational
indoors settling
b. Commercial Filter Makeup air
bullding tmQ
HVAC unit beg Resuspended
particles
é Inflow Filtration
xhaust | Q, (I-E1) fm Ca Qr Et (4m)
Orfm
Lp AnfaR AnVg
Resuspension Gravitational
indoors settling

Figure 8. Transport of soil-derived panicles In the Indoor environment: (a) house with particle
filtration In a heating/cooling unit, and (b) a commercial facility with constant fitration of
fecirculating alr. Unlike the house shown In (a), the farm house is assumed to utilize a Space
hester for heating and so suspended partictes are not removed by filtration.

7


Risk factors(4.9%) Cin/Cs(9.4%) fl CinCs
Mobility
RQ Dose factors
Risk factors

Figure 19, Sources of variance In the estimated cancer risks for the residential exposure scenarto
for Area 13 at 100 y.

41


Kathren, R.L., J.F. McInroy, M.M. Reichert, and M.J. Swint (1988), “Partitioning of 238Pu,
239Pu and 241Am in Skeleton and Liver of U.S. Transuranium Registry Autopsy Cases,”
Health Phys. 54: 181-188.

Kawamura, H. and G. Tanaka (1983), “Actinides Concentrations in Human Tissues,” Health
Phys. 44, Suppl. 1, 451-456,

Kercher, J.R. and L.R. Anspaugh (1991), “Analysis of the Nevada-Applied-Ecology-Group
Model of Transuranic Radionuclide Transport and Dose,” J. Environ. Radioact. 13, 191-216.

Laxen, D.P.H., F. Lindsay, G.M. Raab, R. Hunter, G.S. Fell and M. Fulton (1988), “The
Variability of Lead in Dusts Within the Homes of Young Children,” Environ. Geochem.
Health 10, 3-9.

Layton, D.W. (1993), “Metabolically Consistent Breathing Rates for use in Dose Assessments,”
Health Phys. 64, 23-36.

Lefcoe, N.M. and II. Inculet (1975), “Particulates in Domestic Premises I. Ambient Levels and
Indoor-Outdoor Relationships,” Arch. Environ. Health 3, 565-570.

Leggett, R.W. (1985), “A Model of the Retention, Translocation and Excretion of Systemic Pu,”
Health Phys. 49, 1115-1137.

Lloyd, R.D., G.N. Taylor, W. Angus, F.W. Bruenger, and S.C. Miller (1993), “Bone Cancer
Occurrence among Beagles Given 239Pu as Young Adults,” Health Phys. 64, 45-51.

Market Engineering Corporation (1991), Crystal Ball: A Forecasting and Risk Management
Program for the Macintosh, Ver. 2.0, Market Engineering Corporation, Denver, CO.

Martin, W.E. and S.G. Bloom (1980), “Nevada Applied Ecology Group Model for Estimating
Plutonium Transport and Dose to Man,” in Transuranic Elements in the Environment,
W.C. Hanson, Ed., U.S. Dept. of Energy, DOE/TIC-22800.

McArthur, R.D. (1991), Radionuclides in Surface Soil at the Nevada Test Site, Desert Research
Institute, University of Nevada, Las Vegas, Water Resources Center, Publ. 45077,
DOE/NV/10845-02.

Milford, J.B. and C.I. Davidson (1985), “The Sizes of Particulate Trace Elements in the
Atmosphere—A Review,” J. Air Pollut. Control Assoc. 35, 1249-1260.

Murphy, B.L. and J.E, Yocom (1986), Migration Factors for Particulates Entering the Indoor
Environment, For Presentation at the 79th Annual Meeting of the Air Pollution Control
Association, Minneapolis, MN, June 22-27, 1986, Paper number 86-7.2, pp. 2-15.

Nazaroff, W.W., S.M. Doyle, A.V. Nero and R.G. Sextro (1987), “Potable Water as a Source of
Airborne 222Rn in U.S. Dwellings: A Review and Assessment,” Health Phys. 52, 281-295.

Nicholson, K.W. (1988), “A Review of Particle Resuspension,” Atmos. Environ. 22, 2639-2651.

Ng, Y.C., L.R. Anspaugh and R.T. Cederwall (1990), “ORERP Internal Dose Estimates for
Individuals,” Health Phys. 59, 693-713,

Popplewell, D.S., G.J. Ham, T.E. Johnson, and S.F. Barry (1985), “Plutonium in Autopsy
Tissues in Great Britain, Health Phys. 49, 304-309,

49


5.2. Age-dependent inhalation rates

Respiration provides the oxygen needed to metabolize food nutrients and therefore produce
the energy needed to sustain our various activities. Inhalation rates can be subdivided into two
basic categories, one representing inactive periods of sleep/rest and the other representing the
range of various physical activities carried out each day. Thus, the inhalation rate used for
individuals indoors at home must reflect the respiration rates associated with both active (i.e.,
waking hours) and inactive periods (i.e., sleep/rest); at other indoor/outdoor locations breathing
rates are not assumed to reflect periods of sleep/rest. The breathing rate indoors at home
(designated BRin,h) can be expressed as a weighted-average value of the breathing rate for active
and inactive hours, or

BRin,h = Pr BRe + PaBRa, (13)
where

P, = S/(Ting x 24), (14)

Pg=1-P,, (15)

and

ERinh = breathing rate while indoors at home, m3 d-};
P, = fraction of time indoors at home that is spent at sleep/rest, unitless;
P, = fraction of time indoors at home that is spent at various activities, unitless;
BR, = breathing rate while at rest, m3 d-1;
BR, = breathing rate while active, m3 d~);
S =sleep time, h; and
Tinh = fraction of time spent indoors at home.

The average age-dependent inhalation rate for periods of sleep/rest (i.e., BR;) for males and
females is shown in Figure 12. This rate is based on a methodology presented by Layton (1993)
in which inhalation at rest is determined from the oxygen uptake rate required to support basal
metabolism and a ventilatory equivalent, which is the ratio of breathing rate to oxygen uptake.
During active hours respiration increases to supply additional oxygen to yield metabolic energy
that is equal to the energy expenditures for various activities (i.e., an individual's total daily
energy expenditure less energy expended during sleep/rest). The breathing rate for active hours
can be expressed as a function of the breathing rate at rest or,

BRa=HxBR,, (16)

where

24


1 of 1


reaching the blood stream: 50% will be retained by the skeleton and 30% by the liver. Of the
remainder, most goes to the muscle and skin. The same situation pertains for inhalation, except
that most of the Pu will be swallowed and pass through the gastuintestinal tract without being
absorbed and that which is retained in pulmonary tissue and which will enter the blood has a
removal half-time of 500 days (ICRP, 1986).

Thus, the tissues at risk seem clearly to be the lung, liver, and bone. This is borne out by
many animal studies, with most cancers occurring in lung and bone. The types of bone-related
cancers observed do not include leukemia, but are typically bone sarcomas. This is compatible
with the observation that Pu does concentrate on the bone surfaces, and not in the marrow
(USNAS, 1988). A proup of 26 white male early Pu workers at Los Alamos with known
significant inhalation caposure to Pu have now been followed up for 42 years (Voelz and
Lawrence, 1991). Of the 26, four have died of cancer. Three died of lung cancer, and were
heavy smokers. One died of an osteosarcoma that is most Frobably a result of his exposure
to Pu.

The risk factor to the lung from exposure to alpha particles has received a great deal of
attention lately because of concern about radon. The most recent review of data associated with
radon exposures is that published by the ICRP (1991). On the basis of this teview, the ICRP
concluded that existing studies indicate a lifetime risk of (1-4) x 104 per W  rking Level Month
(WLM) of exposure for members of an average population (the risk for non-smokers is less,
perhaps by a factor of 4). Based on various dosimetry models of the lung and other assumptions,
the calculated dose to the tracheobronchial region per unit of exposure varies from 4-13 mGy
WLM"! (ICRP, 1991). Using the geometric mean of both numbers, we calculate the risk factor
to be 280 x 10-4 Gy-!. Further, if we take the two ranges above 19 represent the 90% range, the
geometric standard deviation (GSD) is calculated to be 1.74. This risk factor is broadly
Consistent with that recently published by Puskin (1992), who calculated a lifetime risk of
2.24 10-4 WLM-! with a 90% range of 1.4 x 104 to 5.7 x 10-4 WLM~!. From data in
Puskin's paper we calculate a geometric standard deviation of 1.53, whereas the equivalent
number for the rangs of (1-4) x 10-4 WLM"! is 1.52.

There have not been as many attempts to derive a risk factor for bone surfaces. The
UNSCEAR (1988) stated that the data would not support the derivation of a risk factor. The
later report of the ICRP (1991), however, supported the use of a value from USNAS (1988,
p. 208), which is derived from a life-table analysis of patients injected with 224Ra. The value is
(133 + 36) x 10-4 Gy-!, which the ICRP corrected for a lethality factor of 0.7. To match our
desire to use lognormal functions, this was converted to a geometric mean of 90 x 104 Gy*t
with a GSD of 1.31. Puskin et al. (1992) have recently pointed out that the original value of
(133 + 36) x 10-4 Gy-! was !...2d on the average skeletal dose, not the dose to the bone surface.
As the dose to the tissue at risk, the bone surface, is 7.5 times higher than the average skeletal
dose, they argue that the risk factor should be divided by 7.5, if it is to be applied to the
calculated dose to the bone surface. This concept has been cautiously endorsed by Bair and
Sinclair (1992). This would result in a risk factor of about 10-3 lethalities Gy-!. However, as
mentioned above, there is one case of bone sarcoma amongst the 26 early Pu workers at Los
Alamos (Voelz and Lawrence, 1991), and ther: seems to be little doubt that this rare cancer is a
radiogenic sarcoma. From the Voelz and Lawrence paper we have estimated that the collective
dose to the done surface of ihe 26 workers was ~20 person—Gy. Thus, the risk factor calculated
from this study would be ~500 x 10-4 Gy7! but the uncertainty (90% range) is from

37


UCRL-ID-112605

Risk Assessment of Soil-Based Exposures to
Piutonium at Experimental Sites Located on the
Nevada Test Site and Adjoining Areas

David W. Layton
Lynn R, Anspaugh
Kenneth T. Bogen

Tore Straume

Sune 1993

Lawrence Livermore National Laboratory
University of California
P.O. Box 808
Livermore, CA 94551-9900

MASTER

S aee cow A ve a ace


determined for the GMX site at the end of a 17-year period. However, subsequent resuspension
studies conducted by Shinn er al. (1986) in the early 1980s showed that the resuspension factor
for contaminated soils at the GMX site had continued to decline to 2.0 x 10-10 mr. They also
reported that the more erodable soils in Plutonium Valley had a resuspension factor of
6.1 x 10-10 m1,

An alternative approach for estimating the amount of a surface contaminant in air is termed
the mass-loading method. With this method, the concentration of a soil contaminant in air is
calculated as the product of the contaminant's concentration in soil and the mass loading of total
suspended particles in air, or

Cc, = C,x TSP, Q)

where Cg is the concentration in air (Bq or |g m73), C, is the concentration in soil (Bq or ug &-!),
and TSP refers to the concentration of particles in the atmosphere (g m~3). Note that the value of
Ca for a contaminant in particulate form can be computed as the product of its concentration in
suspended particles (denoted as Cp) and the mass loading of particles in the atmosphere.
Anspaugh et al. (1974) found that there was good agreement between measured concentrations
of several airborne radionuclides at four sites in the U.S. and concentrations predicted from the
mass-loading method, based on measured values of C, at the sites and an assumed concentration
of suspended particles of 100 1g m-3. An important premise of the mass-loading method is that
the ratio of C, to C, is one. However, this is unlikely to be the case in all instances because the
relationship between the concentrations of a contaminant in the different particle-size fractions of
air and suspendable soil undoubtedly change from site-to-site, depending on the characteristics of
indigenous soils, the physicochemical properties of the contaminant and host soils, ground cover,
and climate. An alternative formulation is to calculate C, as the product of C,, TSP, and an
enhancement factor (EF), that is,

Cy = Cx TSP x EF. @)

Since Ca is equal to C, x TSP, the enhancement factor is then simply the ratio of the
concentrations of a contaminant in suspended particles and in soil. Shinn et al, (1986) reported
that the enhancement factors for Pu aerosols at the GMX site and Plutonium Valley were 0.87
and 1.04, respectively, while the mass loadings of particles were 1.7 x 10-5 and
4,1 x 10°5 g m3, These measurements indicate that the enhancement factor is almost unity for
the two sites.

Neither the resuspension—factor nor the mass-loading approach for estimating resuspension
accounts directly for the wind-induced removal of a soil contaminant at a given site. This loss is
estimated by a resuspension rate or the fraction of contaminant activity or mass that is lost per
unit time. It is equal to the net emission rate from soil (j.e., mass or activity emitted per unit area
and time) divided by the concentration of contaminant on an areal basis (i.e., mass or activity per
unit area). This parameter is estimated indirectly using site-specific data on the changes of
contaminant concentration with height above the ground surface, wind speed, and friction
velocity (Anspaugh et al., 1975). Shinn et al. (1986, 1989) estimated a resuspension rate of
3.9 x 10-11 s-1 for 239.240py in undisturbed soils in Plutonium Valley and 7.9 x 10-13 s-] for

13


“St meie) | ommrreumyemrr{ par so ir0 os

“THE FORGOTTEN GUINEA PIGS”

A REPORT ON HEALTH EFFECTS OF LOW-LEVEL
RADIATION SUSTAINED AS A RESULT OF THE
NUCLEAR WEAPONS TESTING PROGRAM CON-
DUCTED BY THE UNITED STATES GOVERNMENT

REPORT
FEEPAEXD YOR THE USE GF THE
COMMITTEE ON

INTERSTATE AND FOREIGN COMMERCE
UNITED STATES HOUSE OF REPRESENTATIVES:
uD ms “4
SUBCOMMITTEE ON OVERSIGHT AND

INVESTIGATIONS

Ad

AUGUST 1980

U8. GOVERNMENT PRINTING OFFICR
amo ‘WASHINGTON ; 1880

‘Fee sale by the Superintendent of Documents, U.M. Goverament Printing O@ics
‘Washington, D.C. 20603

Fig. 1. Cover of the U.S. Congressional (1980) report on health effects from exposure to
radiation from the Nevada Test Site.


and
Cin = concentration of 239.240Py in indoor air, Bq m-3;
La = loading of 239.240py on floor surfaces, Bq m~2;
fo = fraction of indoor floor dust derived from outdoor soil, unitless;
. Mq = mass loading of particulate matter on floors, g m-?;
Cs = concentration of 239.240Py in soil, Bq 7};
An surface area of floors, m2;
fa = fraction of floor surface that is accessible for walking, playing, etc., unitless;
R = resuspension rate from floors, h-!;
Ach = air-exchange rate, h-};
Ca = concentration of 239:240Pu in outdoor air, Bq m-3;

P = fraction of outdoor 239.240py particles that infiltrate the building shell to indoor air,
unitless;

Vg = gravitational settling velocity, m bh,

V =volume of house (i.e., An times the ceiling height, h), np;

Ef = removal efficiency for particles, unitless;

Qr = recirculation rate of air through heat/cooling unit, mh}; and

fon = fraction of a year that a heating/cooling unit is operating, unitless.

For a commercial facility the steady-state estimate of Cin is (based on Weschler et al., 1983;
Sinclair et al., 1992):

c= —efiAnfaR + Qr(i-EplfmCa ©)
iN Aqvg + QrE¢(1- fm) + Qrfm)’
where

fm = fraction of the air recirculation rate that is
derived from outdoor air, unitless.

Because Ly and Ca are both functions of Cs, Equations 7 and 9 can also be expressed as the
ratios of Cin/Cs (in units of g m-3), or

Cin _ foMgAgfaR + Ach TSP EHP -V

Cs (Agvg + AchV + QrfonE¢]

(for both suburban and farm houses) (10)

and


Breathing rate at rest, mid

2 aot _ L ats L roe a ptt a)
0 10 20 30 40 50 60 70 80
Age, years

Figure 12. Average Inhalation rate at rest for males and females as a function of age (derived from
Layton, 1993).

"1 = (24M ~ S)(24~S), (17)

and
H = breathing-rate—multiplier used to calculate BRg as a function of BR,, unitless; and

M = ratio of total energy expended each day to an individual's basal metabolic rate (BMR),
unitless.

In this formulation H represents the factor that BR, must be increased by to meet the
increased metabolic requirements associated with waking hours. The two age-dependent
parameters used to calculate H are sleep time (S) and the ratio of total energy intake to BMR
(also referred to as the metabolic equivalent, denoted here as M). As sleep time decreases and M
increases, the amount of energy expended during active hours increases with respect to the
resting inhalation rate, and there is a commensurate increase in the value of H. Figure 13 shows
the metabolic equivalents of the total energy expenditures for different ages (from Layton, 1993),

25


Introduction to Section II and
Overview of Dose Reconstruction:
Lessons Learned from Studies in the U.S.

Lynn R. Anspaugh’
Dose Reconstruction Program
Lawrence Livermore National Laboratory
Livermore, CA 94550

INTRODUCTION

Table 1 outlines our schedule for this part of the session. I am going to give an overview
of dose reconstruction with some emphasis on the lessons that we have leamed from work in the
United States. Then, Dr. Marina Degteva will talk about dose reconstruction for the cohort of
people living near the Techa River downstream from the Mayak Production Association, which
was the first nuclear materials production facility in the former Soviet Union. We will then

conclude with an open question/answer/discussion session on dose reconstruction.

As mentioned, the purpose of this presentation is to provide an overview of dose
Teconstruction with an emphasis on the lessons learned from work in the United States. Several
major dose reconstructions have been undertaken in the United States, particularly in reference to
Department of Energy (DOE) facilities. Some of these activities have now been completed and
these are indicated in the upper part of Table 2. The first major activity took place at the Nevada
Test Site (NTS), where researchers have considered several different specific populations. The
activities began with an analysis of hypothetical individuals, which was followed by an analysis of
the collective dose to all exposed individuals within the surrounding region. Later, the University
of Utah undertook some specific epidemiologic studies and calculated doses to specific
individuals. The Hanford Environmental Dose Reconstruction Study has completed its results for
hypothetical individuals. The Hanford researchers did not report collective dose.

* Present address: Radiobiology Division, University of Utah, Salt Lake City, UT 84112


soils are expected to provide a reasonable indication of the levels of 239.240pu in soils. FIDLER
surveys were conducted across uniform grids for all sites, except GMX where measurements
were made on a radial grid. Isopleths of gamma activity (counts per minute of 60-keV gamma
rays) based on the survey data were then used to designate strata for subsequent soil sampling to
determine concentrations of 239.240py, Figures 3 through 6 present the isopleths of 241Am
activity and associated strata for the various areas. The sampling locations within the irregular
boundaries of each stratum were determined using a stochastic procedure in which the horizontal
and vertical distances on a grid over a site were selected randomly to define a sampling
coordinate (i.e., the intersection of the two random distances). The number of soil samples in the
strata varied depending on the areal extent of the strata and the expected sample variability for a
given stratum. However, at least 10 soil samples were taken in each stratum, regardless of its
size. Soil samples were taken to a depth of 5 cm and later 10-gram aliquots of soil were
homogenized by ball-milling prior to analysis. Other soil samples were taken to depths of 25 cm
in order to evaluate the distribution of 239.240py concentrations with depth. Gilbert et al. (1975)
reported that between 68 and 97% of the total plutonium in soil to 25 cm was present in the top
Scm. Thus, the inventories based on the 5-cm depth may slightly underestimate the total amount
of 239,240Py present in the near-surface soils. In addition, the estimated inventories exclude an
unknown amount of Pu present in the soils and other shot-related debris that were put in trenches
or otherwise covered with clean soil during cleanup operations.

Table 1 summarizes the results of the analyses of the inventories of 239.240Pu in soils at the
shot sites. An estimated 5.1 TBq of 239.240py is distributed in the upper 5-cm soils at the sites.
The location with the highest inventory is Area 13 (33% of the total), however, the highest levels
of contamination of surface soils occur at Area 11 (Plutonium Valley), where as much as
3070 Bq g- is present in one of the survey strata. The concentrations cf 239.240pu in the soils of
each of the sites were determined by dividing the estimated inventory for a site by the mass of
contaminated soil (based on a density of 1.5 g/cm3). Appendix A contains the strata~specific
data on 239249py at each of the shot sites. In later work, scientists associated with the
Radionuclide Inventory and Distribution Program (RIDP), which was designed to characterize
the spatial distribution and inventory of selected radionuclides at the NTS, made measurements
of 41am at the GMX and Plutonium Valley sites using in-situ gamma spectrometry (McArthur,
1991), Those measurements were converted to estimates of 239.240Py activity in soil and then
estimates were made of the 239.240Py inventories at those two sites. The resulting RIDP
estimates were 0.052 and 1.1 TBq (1.4 and 29 Ci) for GMX and Plutonium Valley, compared
with the NAEG estimates of 0.056 and 1.3 TBq (1.5 and 36 Ci), indicating close agreement.

3. Conceptualization of Potential Exposure
Pathways and Screening Analyses

From an exposure-assessment standpoint, one of the most important determinants affecting
the nature and magnitude of future contacts with residual contaminants in soils is the type of land
use that eventually occurs at a given location. We have identified four plausible land uses for the
desert region where the test shots were located: (1) cattle grazing, (2) resident ranching,
(3) residential housing, and (4) commercial development. Grazing is currently practiced


Identification of toxic responses that are route specific

~»>/ potential exposure pathways

-| response relationships and/

Process

Characterization of
residual Pu in soll

Conceptualization of

and screening analyses

Prediction of Pu transport
In the desert environment

Estimation of exposures
to Pu for alternative
land-use scenarios

Determination of
organ-specific doses

Assessment of dose-

or risk factors for organs

Characterization of
cancer risks

Content

Obtain and analyze data on the
conceiitrations and inventories
of Pu in soil.

Identify potential exposure pathways
and conduct screening analyses to
identify salient pathways.

Determine the time-dependent
concentrations of Pu in air and soil.

Based on the screening analyses

and evaluation of potential land uses,
estimate exposures to Pu for alternative
scenarios and routes of exposure.

Determine doses to critical organs
addressed in the dose-response
assessment.

Review pertinent animal studies

and epidemiological data to construct
dose-response relationships or risk
factors for specific organs.

The final step involves the integration
of all of the previous steps to estimate
potential health risks and associated
uncertainties.

Figure 2. Methodology used to assess the risks associated with residual 239,240py in solts at the
various shot sites.


4

08 preesrreerpreeers

Suburban

0.4

Fy i oe Conon meee eee

we sees owes

Probability

0.2

0 10 20 30 40 50 60 70

08 prepreg prec prerpeypee

Farm 1

Probability

0 poet tid

Oo 8610 20 30 40 50 60 70
Years of residence

Figure 15. Mobility of persons fiving in rural suburban residences and farms (derived from U.S.
Bureau of the Census, 1991}.

28


12

Haskell, E.H., Bailiff, .K., Kenner, G.H., Kaipa, P.L., and Wrenn, M.E. Thermoluminescence
Measurements of Gamma-Ray Doses Attributable to Fallout from the Nevada Test Site
Using Building Bricks as Natural Dosimeters. Health Phys. 66: 380-391 (1994).

Haskell, E.H., Kenner, G.H., and Hayes, RB. Electron Paramagnetic Resonance Dosimetry of
Dentine Following Removal of Organic Material. Health Phys. 68: 579-584 (1995),

Heeb, C.M. Radionuclide Releases to the Atmosphere from Hanford Operations, 1944-1972.
Pacific Northwest Laboratories, Richland, WA, PNWD-2222 HEDR, 1994.

Henderson, R.W., and Smale, R.F. External Exposure Estimates for Individuals Near the Nevada
Test Site. Health Phys. 59: 715~721 (1990).

Henderson, R.W., and Smale, RF. Summary of Collective Dose from External Exposure.

Los Alamos National Laboratory, Los Alamos, NM, Report submitted to the Coordination
and Information Center, Las Vegas, 1992.

Hicks, H.G. Additional Calculations of Radionuclide Production Following Nuclear Explosions
and Pu Isotopic Ratios for Nevada Test Site Events. Health Phys. 59: 515-523 (1990).

IAEA. The Radiological Impact of Radionuclides Dispersed on a Regional and Global Scale:
Methods for Assessment and their Application. International Atomic Energy Agency,
Vienna, Technical Reports Series No. 250, 1985.

Kerber, R.A., Till, J.E., Simon, S.L., Lyon, J.L., Thomas, D.C., Preston-Martin, S., Rallison,
MLL., Lloyd, R_D., and Stevens, W. A Cohort Study of Thyroid Disease in Relation to
Fallout from Nuclear Weapons Testing. J. Am. Med. Assoc. 270; 2076-2082 (1993).

Kirchner, T.B., Whicker, F.W., Anspaugh, L.R., and Ng, Y.C. Estimating Internal Dose Due to
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Nuclear Test Fallout. National Technical Information Services, Springfield, VA, TID-
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Nevada Test Site Detonations. New York: U.S. Department of Energy Environmental
Measurements Laboratory, EML—400, 1981.


3.15 x 10-5 to 3.15 x 10-4 y-! (encompassing the range of resuspension rates discussed earlier)
and with a mode at 1.7 x 104 yl. We were only able to obtain two measurements of TSP at the
NTS (see earlier discussion), and so we completed a statistical analysis of TSP levels in 20 rural
locations in the U.S. reported by Shah et al. (1986) as a way of characterizing the variability in
the particulate mass loadings at the shot locations. The geometric mean of the TSP data was
2.8 x 10-5 g m-3 with a GSD of 1.6. The value of EH is set at 1, based on the experimental
studies noted above. :

Since 239240py at the shot sites could pose a radiological health hazard for thousands of
years if not remediated, we computed values of Cs for time periods of 100, 10,000, 50,000, and
100,000 y to cover the period of time when residual 239,240py in soils are likely to constitute a
health hazard. At 100 y the 5, 50, and 95% cumulative percentile values of C, at Area 13 are 5.0,
5.4, and 5.7 Bq g-!, respectively, based on 10,000 random samples (using the Latin hypercube
sampling option) of the distributions for Cs and Ar. However, at 100,000 y the percentile values
are 1.8 x 10-6, 1.8 x 10-5, and 1.5 x 10-4 Bg g-}. These results indicate that as ¢ increases, the
variation in C; becomes dominated by the variation in Ar.

4.2. Concentrations of Pu in Indoor Air

The concentrations of certain airbome contaminants indoors will be less than those outdoors,
but for others the concentrations indoors may be higher. A primary contributing factor is the
source of each specific contaminant. For those contaminants whose primary source is outdoors,
as would generally be the case for Pu, the concentration indoors will generally be less than that
outdoors. This is due, among other things, to the filtering effect of building shells (see
Engelmann, 1992). To determine the levels of 239240Py in the indoor environment of a house
(suburban residence and farmhouse) and a commercial facility as a function of the contamination
of soil and outdoor air, we have implemented two simple compartmental models that predict the
steady-state concentrations of 239.240py in the indoor air. As shown in Figure 8a, the transport
processes for a house include the infiltration of contaminated outdoor air, exfiltration of building
air, resuspension of tracked-in soil particles from floors to indoor air, the gravitational settling of
airborne particles to floors, and the filtration of particles by a heating/cooling system, For a
commercial facility (Figure 8b) with an air-handling system that exerts a positive air pressure
indoors, there is no significant infiltration of outdoor air (and associated airborne particles).
Instead, outdoor air is continuously introduced to the air-handling system and exhausted from the
building. The concentration of 239.240py in the indoor air of a house under steady-state
conditions (and with no intemal sources of 239.240py other than resuspension) can be estimated
from

Lg AgfaR + AchCaPV

Cin = 7)
" (Anvg + AchV + QrfonEF) ‘
where
La =fo Ma Cs (8)


airborne 239.240py) for the suburban house, farmhouse, and commercial building were 0.24, 0.32,
and 0.19, respectively.

5. Human Factors Describing Exposure Scenarios

To quantify inhalation exposures to 239,240Pu from soil-derived aerosols in indoor and
outdoor air and the cumulative doses resulting from those exposures, it is necessary to
characterize the following parameters: (1) the age-dependent breathing rates for indoor and
outdoor locations, (2) the fractions of time spent indoors and outdoors at residential and
commercial sites where soil contamination occurs (i.e., the exposure frequency), and (3) the
length of time spent at each location (i.e., the exposure duration). The locational aspects of
human activities are important because the concentrations of 239240py differ between indoor and
outdoor air. Moreover, the amount of time people spend at a given location depends on its
function (e.g., a residence or place of work). Inhalation rates change from location to location as
a function of the physical activities carried out in the different locations. The four basic locations
that people occupy are shown in Figure 9. For each location type we will now characterize
exposure factors that will be used in the dose assessments for alternative land-use scenarios at the
shot sites.

5.1. Activity Patterns

Important exposure factors that must be considered when addressing contacts with soil
contaminants at a given location (i.e., nonubiquitous soil contaminants) is the amount of time an
individual spends at indoor and outdoor locations at the site and the kinds of activities performed
while there. Both of these lifestyle factors changes with age. We have characterized the age-
dependent changes in the amount of time an individual spends at indoor and outdoor locations at
home using the results of activity surveys of children (< 12 y) and adults (2 12 y) in California
(Wiley, 1991; Wilcy er af., 1991), The random sample of children included 1200 individuals,
while the adult sample included 1,762 individuals. Figure 10 shows the fractions of each day
that were spent indoors and outdoors at home. As might be expected, there are U-shaped
functional relationships between age and the fractions of time spent indoors and outdoors at
home. The fraction of time indoors decreases from birth to about age 20 y and then gradually
increases into the retirement years. By comparison, the amount of time outdoors at home
averages less than 5% of a day for those under 10 y, decreasing to around 2% in middle age, and
then increasing again in the retirement years. The amount of sleep time is presented in
Figure 11. Children under age 10 sleep more than 10 h per day, and as individuals reach middle
age, they sleep an average of eight hours per day. Small increases in steep time occur after age
50 y. Unfortunately, we were unable to obtain similar age-dependent activity-pattern data for
residents of farms. If such data were available, we would expect that farm residents would spend
much more time outdoors adjacent to a farm house. We will examine the effect that increased
time outdoors has on our dose estimates for the resident farmer in the section dealing with dose
estimates.

21


14

Reiss, E. Statement of Eric Reiss, M.D., Associate Professor of Medicine and Preventive
Medicine, Washington University School of Medicine; Director, Irene Walter Johnson
Institute of Rehabilitation. In: U.S. Congress (1963), Part 2, pp. 601-672.

Simon, S.L., Till, JE., Lloyd, RD., Kerber, R.L., Thomas, D.C., Preston-Martin, S., Lyon, J.L.,
and Stevens, W. The Utah Leukemia Case-Control Study: Dosimetry Methodology and
Results. Health Phys. 68: 460-471 (1995).

Stevens, W., Thomas, D.C., Lyon, J.L., Till, J.E., Kerber, RA., Simon, S.L., Lloyd, R.D.,
Elghany, N.A., and Preston-Martin, S. Leukemia in Utah and Radioactive Fallout from the
Nevada Test Site. J. Am. Med. Assoc. 264: 585-591 (1990).

Stevenson, K.A., and Hardy, E.P. Estimate of Excess Uranium in Surface Soil Surrounding the
Feed Materials Production Center Using a Requalified Data Base. Health Phys. 65: 283-
287 (1993).

Straume, T., Marchetti, A.A., Anspaugh, L.R., Khrouch, V.T., Gavrilin, Yu.l., Shinkarev, S.M.,
Drozdovitch, V.V., Ulanovsky, A.V., Komeev, S.V., Brekeshev, M.K., Leonov, E.S.,
Voigt, G., Panchenko, S.V., and Minenko, V.F. The Feasibility of Using !29] to
Reconstruct !3!1I Deposition from the Chemobyl Reactor Accident. Health Phys. 71, 733-
740 (1996).

Till, JE., Simon, S.L., Kerber, R., Lloyd, R.D., Stevens, W., Thomas, D.C., Lyon, J.L., and
Preston-Martin, S. The Utah Thyroid-Cohort Study: Analysis of the Dosimetry Results.
Health Phys. 68: 472-483 (1995).

UN. Scientific Committee on the Effects of Atomic Radiation. Exposures Resulting from
Nuclear Explosions. Annex E in: /onizing Radiation: Sources and Biological Effects.
United Nations Scientific Committee on the Effects of Atomic Radiation. 1982 Report to
the General Assembly, with Annexes. United Nations, New York, Sales No. E.82.[X.8,
1982, pp. 211-248.

UN. Scientific Committee on the Effects of Atomic Radiation. Exposures from the Chernobyl
Accident. Annex D in: Sources, Effects and Risks of lonizing Radiation. United Nations
Scientific Committee on the Effects of Atomic Radiation 1988 Report to the General
Assembly, with Annexes, United Nations, New York, Sales No. E.88.[X.7, 1988, pp. 309-
374.


DISCLAIMER

This document was prepared as an account of work sponsored by an agency of
the United States Government. Neither the United States Government nor the
University of California nor any of their employees, makes any warranty, express
or implied, or assumes any legal liability or responsibility for the accuracy,
completeness, or usefulness of any information, apparatus, product. or process
disclosed, or represents that its use would not infringe privately owned rights.
Reference herein to any specific commercial product, process, or service by trade
name, trademark, manufacturer, or otherwise, does not necessarily constitute or
ply its endorsement, recommendation, or favoring by the United States
Government or the University of California. The views and opinions of authors
expressed herein do not necessarily state or reflect those of the United States
Government or the University of California, and shall not be used for advertising
or product endorsement purposes.



ir
All

Association for Informati 1d Image Management
rT ne. ite 1400

1

2 3 4 5 6 7 8 © 1 1 12 13 14 15 mm
4

1

Centimeter

Inches

MANUFACTURED TO AIIM STANDARDS

BY APPLIED IMAGE, INC.


26 x 10-4 Gy to 2120 x 16-4 Gy-!. While the associated uncertainty is large, one would have a
Cifficult time arguing that the actual risk from chronic exposure is near zero or far less than the
value of 90 x 10-4 derived above. Also, the surviving 22 individuals are now elderly and
therefore it is not anticipated that they will accumulate a greatly increased collective dose. At the
same time, however, it is quite possible that yet more bone sarcomas will be found.

Lloyd er al. (1993) have recently published the results of a lifespan study of beagles injected
with Pu. They extrapolate to man a risk factor of 548 x 10-4 Gy! for the average dose to the
skeleton. This is equivalent to a risk factor of 70 x 104 Gy"! for dose to the bone surfaces.
Thus, we have elected to use the value of 90 x 104 Gy-!, but it seems clear that the GSD value
of 1.31 is too small. This is especially true with consideration of the reported wide variation in
the amount of Pu that does end up in the skeleton (Kathren et al., 1988). Thus, the GSD (1.74)
for the lung-risk factor has been used as a more likely default value for the GSD for the bone-
surface-tisk factor.

Even fewer data are avaiiable for the induction of liver cancer by radiation. The only solid
body of evidence concerns the many individuals who were injected with 232Th in the form of
“Thorotrast.” Again, the ICRP (1991) refers to the analysis of lifetime excess risk done by the
USNAS (1988). The calculated risk coefficient for three major studies varies from
(260-300) x 10-4 Gy~). For our calculations, we will use the value of 280 x 104 Gy-! and also
use the GSD of 1.74 for the lung-risk factor as the default value for the GSD for the liver-risk
factor.

7. Predicted Cancer Risks and Associated
Uncertainties

The cancer risks for individuals inhabiting a suburban residence, living in a farm house, or
working at a commercial facility at the shot locations were calculated for periods of 100, 50,000,
and 100,000 y in order to evaluate the long-term risks associated with the possible loss of
institutional controls at the NTS. As a cautionary note, we stress that our risk estimates for such
lengthy time frames are predicated on the existence of exposure scenarios that are consistent with
our previous characterizations (i.e., activity patterns, housing properties, etc.). Results of the
site-specific risk assessments are shown in Table 5 as the 5, 50, and 95% percentiles of
incremental cancer risk from the cumulative probability distributions generated by the Monte-
Carlo simulations. The resident farmer is shown to have the highest risk of the three exposure
scenarios. Principal factors contributing to that risk are long tenure and higher concentrations of
239,240Py in indoor air due to the lack of indoor filtration afforded by a central heating/cooling
unit. At the median risk level, the resident farmer's risk is more than a factor of three higher than
the suburban resident and about a factor of ten higher than the worker at a commercial facility.
The resident farmer's estimated tisk, though, could be a factor of five to ten higher if additional
assumptions were made regarding time spent outside adjacent to a farmhouse and an increased
inhalation rate associated with physical labor. These results demonstrate that exposure scenarios
incorporating lifestyle factors such as mobility and indoor/outdoor relationships for soil-derived
contaminants can result in significant differences in the levels of cancer risk predicted for
individuals whose lifestyles match the reference scenarios.

38


3

Table 8. Several specific issues and events clarified the need for a dose-reconstruction study for
the persons living downwind of the Nevada Test Site during the 1950s and 1960s.

Concern about the implications of a reported increase in the incidence of leukemia in atomic
veterans, which was first reported in 1976 by the Centers for Disease Control and Prevention
(Caldwell et al. 1980)
Several hundred personal injury claims from residents in Nevada, Utah, and Arizona
Congressional Hearings (1979-1980) (e.g., U.S. Congress 1979, 1980)

e “Atomic Veterans”

¢ Forgotten Guinea Pigs”
The Lyon et al. (1979) epidemiologic study implied an increase in leukemia in “high fallout”
locations.
Concem about the correctness of past analyses of external doses and rekindling of concern
about the unquantified doses from internal exposure (from inhalation and ingestion)
Two legal cases

¢ Reopening of Bullock et al. vs. U.S. (sheep deaths)

«Allen et al. vs. U.S. (human cancer)



UCRL-JC-126357
PREPRINT

Introduction to Section II and Overview of
Dose Reconstruction:
Lessons Learned from Studies in the U.S.

LR. Anspaugh

This paper was prepared for submittal to the
Symposium on Assessing Health and Environmental Risks from
Long-Term Radiation Contamination in Chelyabinsk, Russia
at the 1996 American Association for the Advancement of Science
and Science Innovation Exposition
Baltimore, MD
February 8-13, 1996

January 1997

‘Thisis a preprint of a paper intended for publication in e journal or proceedings.
Sincechanges may be made before publication, this preprint is made available with
theanderstanding that it will not be cited or reproduced without the permission of the
anthor.



Area 13 ‘Net counts per minute of 24¢Am

isopleth 1 mlm 0-1,000

2 [3 1,000-5,000
3 E3 5,000-10,000
4 ml 10,000-25,000
5 tm 25,000-50,000
Outer fence 6 E> >50,000

2

Inner fence

Seale

—
Feet

Area 5 - GMX Site

z

Net counts per minute of 241Am

Strata 1 = 500-5000

2 mmm 5000-25,000
3 &=) 25,000-50,000
4 wm >50,000

5 t= 0-50,000

Ground zero

Figure 5, Areal extent of the contamination of 241m in solls at Area 13 and Area 5 (GMX). The
measured 241m is used as a surrogate for 239.240py (adapted from Gilbert ef al., 1975).


Technical Information Department » Lawrence Livermore National Laboratory
University of California + Livermore, California 94551



Introduction

Residual plutonium (Pu) in surficial soils at the Nevada Test Site (NTS) is the result of the
above-ground testing of nuclear weapons and special experiments involving the detonation of
plutonium-bearing devices. The latter experiments were conducted for four basic reasons
(Stannard, 1988): (1) to study the behavior of Pu as it was being explosively compréssed, (2) to
ensure that the accidental detonation of the chemical explosive in a production weapon would not
produce a nuclear yield, (3) to evaluate the ability of personnel to handle large-scale Pu dispersal
accidents, and (4) to develop criteria for transportation and storage of weapons. The first set of
experiments were called “hydrodynamic” or “equation of state” tests. Twenty-two such tests
were conducted above ground at the GMX location in Area 5 (see Fig. 1) between December
1954 and February 1956. Many tests were conducted to determine whether nuclear yields could
be produced from accidental detonations. The largest ones carried out on the surface were
Project 56 in Area 11 (now known as Plutonium Valley); here four devices were detonated in
1956, one with a slight nuclear yield. The final large, surface experiment was conducted as
Project 57 in Area 13, just off the NTS. Experiments performed to evaluate cleanup and
weapons-handling issues consisted of Operation Roller Coaster, Double Tracks, and Ciean Slate
I, If, and IIL. The locations of these detonations off of the NTS are also indicated in Figure 1.

All of the sites where these tests were conducted were contaminated with Pu. At the present
time, these sites do not pose a health threat to either workers or the general public because they
are under active institutional control. Specifically, inadvertent human intrusion onto the test sites
(referred to here as “shot sites”) is deterred by the use of fences and warnings, coupled with the
fact that the shot sites are within test ranges, all of which have restricted access. However,
because the half-lives of 239Pu and 240Py are 24,110 and 6,560 y, respectively, the residual
contamination could indeed pose a long-term health hazard if it were not remediated and
industrial, agricultural, or residential land uses emerged following a loss of institutional control.
In order to develop relevant cleanup-options for the test sites, it is important to assess the nature
and magnitude of the potential health risks associated with naive uses of the affected lands,
Accordingly, the following risk analysis examines the distribution of plutonium in soils at these
sites, potential exposures and resulting doses, and finally, incremental cancer risks for those
potentially exposed.

1. Methodology Overview

The basic methodology for estimating the potential health risks of residual 239.240py in the
soils of the NTS is shown in Figure 2. The initial task involves the characterization of the
concentrations of 239.240Py in surface soils in the vicinity of each test shot. That effort is
followed by a conceptualization of how an individual could come in contact with the residual
plutonium under various land-use scenarios following loss of institutional control of
unremediated lands. Knowledge of route-specific toxicity is also needed to identify potentially
important exposure pathways, and hence we have shown a feedback loop between the dose-
response assessment and the conceptualization of potentially important exposure pathways.


NOTE: CUMULATIVE ESTIMATED EXPOSURE
{W) FOR ALL NEVADA TESTS
THROUGH tesa,

’

FALLOUT PATTERN-ALL TESTS

Fig. 2. Isopleths of estimated individual extemal gamma exposure from the nuclear tests
conducted at the Nevada Test Site (from Friesen 1985).


10

are not yet available. In general it seems likely that further dose-reconstruction and epidemiologic
studies at U.S. sites will not contribute to scientific knowledge conceming the dose-effect
selationship of cancer induction by radiation exposure.

If the problem of dose-effect is to be addressed, particularly as it pertains to doses
delivered at low-to-moderate dose rates, meaningful results are much more likely to follow from
studies conducted in Russia at sites related to the production of nuclear materials and from studies
in Belarus, Ukraine, and Russia related to the Chernobyl accident.

ACKNOWLEDGEMENTS

This work was performed under the auspices of the U.S. Department of Energy at the
Lawrence Livermore National Laboratory under Contract No. 48-Eng-7405.

REFERENCES

Anspaugh, L.R. Technical Basis for Dose Reconstruction. In Proc. Of the 31* Annual Meeting
of the National Council on Radiation Protection and Measurements, Arlington, VA,
April 12-13, 1995 (in press).

Anspaugh, L.R., and J.1. Daniels, J.I. Bases for Secondary Standards for Residual Radionuclides
in Soil and Some Recommendations for Cost-Effective Operational Implementation. Health
Phys. 70, 722-734 (1996).

Anspaugh, L.R., Ricker, Y.E., Black, $.C., Grossman, R.F., Wheeler, D.W., Church, B.W., and
Quinn, V.E. Historical Estimates of External y Exposure and Collective External y
Exposure from Testing at the Nevada Test Site. Il. Test Series After Hardtack Il, 1958,
and Summary. Health Phys. 59: 525-532 (1990).


Hours 3 sleep per day

4
OQ Gee

0 10 20 30 40 59 60 70
Age, years

Figure 11. Hours of sleep as a function of age (data from Wiley, 1991; Wiley ef a/., 1991).

The amount of time spent at a commercial facility depends in part on an individual's
employment status (i.e., full- or part-time worker) and the nature of the facility (e.g., retail,
office, manufacturing, etc.). Because our commercial land-use scenario is not meant to represent
any particular kind of commercial operation, we will use labor statistics for the employed civilian
work force in the U.S. to characterize the time spent at an unspecified work location.
Employment data (U.S. Bureau of the Census, 1991) show that the work week for wage and
salary workers employed in nonagricultural work has averaged between 38 and 39 h since 1970.
Expressed on the basis of a seven-day week, this translates to about 23% of an average day (or
on an annual basis, 84 d y-!). We were uncble to obtain data on the amounts of time that
workers spent outdoors at office, retail, and manufacturing locations. For tne purposes of our
exposure assessment, we have assumed simply that workers in the arid Nevada climate would
spend 0.5 h d-! outdoors at their place of employment (or 5.4 d y~!, based on a 52-week
v-orking year). One additional consideration in characterizing the worker population is the age
span of those employed. Although individuals can enter the workforce at age 16 y, we assume
that full-time employment begins at age 19 y and continues until age 64 y. Workers aged 65 y
and older and younger than 19 y accounted for less than 10% of the total civilian workforce in
1989 (U.S. Bureau of the Census, 1991).

23


Table 5. Results of Monte-Carlo simulations of the cancer risks associated with three
different land uses at three shot sites at the NTS.

Cel

Cumulative cancer risk
Site Year Land use 5% 50% 95%
Area 13 100s Residential 7x18 7x10-7 7x10
Farm 1x107 2x10 2x 105
Commercial 2x18 3x10-7 3x 10-6
10000 _— Residential 8x10% gx10°8 1x 106
Farm 2x108 3x10-7 3x 10-6
Commercial 5x10 sx10-8 5x10-7
50000 = Residential 6xiol extol? 6x10?
Farm 2xigle 2x10? 2x10°8
Commercial 9x10 9x10-!0 3x 109
100000 —- Residential 3x 10-12 3x10 7x 1071!
Farm 4x 10-12 axigl Bxiol
_ Commercial 2x19°!2 2xigh 5x 10-11
Clean Slate If 100 —- Residential 21077 2x106 2x 10-5
Farm 5x 10-7 7x10 5x 10-5
Commercial 8x 10-8 1x 196 1x10°5
10900 _—Residential 3x 10-8 3x10-7 axir6
Farm 9x 10-8 9x 10-7 9x 10-6
Commercial 2x 10% 2x 10-7 2x10
50,000 ~—sResidential 2x 19-10 2x10-9 2x08
Farm Bx 1g l0 ax to? 6x10°8
Commercial 3x 19-10 3x 109 9x 10-9
100000 _—- Residential 1x10)! 1x 19-10 2x 19-10
Farm 2x10}! 2x 190710 4x1g-l0
Commercial 8x tol2 exig tl 2x19710
Area 11D site 100 _—Residential 2x 10-6 2x10°5 2x104
Farm 4x 10-6 6x 10-5 5x10-4
Commercial 6x 10-7 8x 10-6 9x 10-5
10000 ‘Residential 2x107 3x10% 3x10-5
Farm 7x17 8x10% 8x10-5
Commercial 1x107 1x10 1x10°5
50,000 ‘Residential 2x109 2x10°8 2x10-7
Farm 1x108 1x107 5x10-7
Commercial 3x 10-9 3x10°8 8x10-8
100,000 ~—s Residential 1x 19710 1x10% 2«10%
Farm 3x 10710 3x10 5x109
Commercial 2x19! 7x 19710 1x109

39


t Area 11 — Plutonium Valley
N : D Site

Feet
ASite
Net counts per minute of 24tAm
Strata 1c <5000
2,3,4,5 mmm >5000
6 5000-10,000
ASite mmm 235U

Figure 6. Areal extent of the contamination of 241Am in soils at Area 11 (Plutonium Valley). The
measured 247am Is used as a surrogate for 239,240py (adapted from Gilbert et al., 1875).

Table 1. Summary of the 25°240pu inventories at the shot sites and associated concentrations
in soil (from Gilbert, 1977).

Inventory Concentration in soil
Site Area ha TBq +SE Bq a7! +SE
Area 13 400 17 0.33 5.6 1
Area 5 (GMX) R 0.056 0.007 59 0.75
Double Tracks 18 0.13 0.035 9.9 26
Clean Slate I 18 0.16 004 12 33
Clean Slate II a7 0.63 . 0.137 18 39
Clean Slate 11 170 14 0.20 10 15
Area 117
B site 18 0.23 0.041 175 31
C site 2.6 0.29 0.063 149 32
Daite 53 0.63 012 158 30
C/D Overlap 62 0.028 0.012 6 25
—

SE EA OO LE RE
® The area encompassing the region where there was less than 8,000 cpm of 241Am contained an additional
9.17 Bg of Bibby,


Inhalation Dose model

NY |
i res * an, >| Target organs
Loe ;
. ‘| Organs * kidney H
' * liver '
Accidental : iungs resp. ta . bone ,
ingestion ‘| « lymph nodes *totalbody = fi
Wash ! ;
Peel * Gi model
Muscle [~~ rooseseesesessses
H Gl tract :
Beef cattle weer a - Dairy cattle
. liver ° Liver Milk e milk
Cultivated Desert
vegetables vegetation Alfalfa
Soil (C,) . J

Figure 7. Conceptual diagram of the exposure pathways for ingesting and inhaling 2282404 in
the desert environment of the NTS (adapted from Kercher and Anspaugh, 1991).

a0)


13

Likhtarev, L.A, Sobolev, B.G., Kairo, I.A., Tronko, N.D., Bogdanova, T.I., Oleinic, V.A.,
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